Asian Longhorned Beetle

Eradication Program

Final Programmatic 

Environmental Impact 

Statement—September 2015

 

United States 

Department of 

Agriculture 

Marketing and 

Regulatory 

Programs 

Animal and 

Plant Health 

Inspection 

Service 
 

 
Asian Longhorned Beetle 

Eradication Program

Final Programmatic Environmental 

Impact Statement—September 2015

Agency Contact: 

Jim E. Warren, Ph.D. 

Ecologist 

Policy and Program Development 

Animal and Plant Health Inspection Service 

U.S. Department of Agriculture 

4700 River Road, Unit 149 

Riverdale, MD 20737–1238 

 
 

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Table of Contents i.

Table of Contents 

Executive Summary................................................................................. v 

I. Purpose of and Need for Action ............................................................. 1 

 A. Background ...................................................................................... 2 

1. Description of ALB ..................................................................... 2 

2. Economic Impacts of ALB in the United States .......................... 7 

3. Existing and Projected Ecological Impact in the United States 

and Elsewhere ........................................................................... 9 

 B. Public Involvement ......................................................................... 11 

1. Prior EAs Published by the Agency and Results from 

 Public Comments .................................................................... 11 

2. Past Agency Outreach Efforts on ALB Issues .......................... 11 

3. Locations of Addressed Issues Raised by Stakeholders ......... 11 

C. Decision Framework ...................................................................... 15 

D. Scope of this Document and NEPA Requirements ....................... 17 

E. Consultations ................................................................................. 17 

II. Alternatives .......................................................................................... 19 

A. Quarantine ..................................................................................... 19 

B. Survey ........................................................................................... 20 

1. Level 1 Survey (Core Survey) ................................................. 21 

2. Level 2 Survey (Buffer) ............................................................ 21 

3. Level 3 Survey (High Risk Survey) .......................................... 22 

4. Secondary Survey ................................................................... 23 

C. Alternatives Evaluated in this EIS ................................................. 23 

1. No Action ................................................................................. 25 

2. Removal of Infested Trees ...................................................... 26 

3. Full Host Removal ................................................................... 27 

4. Insecticide Treatment .............................................................. 28 

5. An Integrated Approach (Preferred Alternative) ...................... 29 

D. Alternatives Considered but Not Included in this EIS .................... 30 

1. Use of Other Chemical Control Agents ................................... 30 

2. Other Integrated Pest Management / Suppression 

Strategies ................................................................................ 30 

III. Affected Environment .......................................................................... 33 

A. Affected Forest Communities ........................................................ 33 

B. Affected Water Resources ............................................................ 34 

1. Water Quality ........................................................................... 35 

2. Interactions of Soils with Water Quality ................................... 37 

C. Affected Soil .................................................................................. 37 

1. Soil Types of the United States ............................................... 38 

2. ALB-Host Species and Soil Quality ......................................... 38 

3. Soil Erosion ............................................................................. 39 

D. Air .................................................................................................. 40 

E. Affected Biological Organisms ...................................................... 43 

F. Affected Economic, Social, and Cultural Factors .......................... 46 
ii. Table of Contents 

1. Market Factors ......................................................................... 46 

2. Timber ..................................................................................... 47 

3. Maple Syrup ............................................................................ 50 

4. Tree Nurseries ......................................................................... 50 

5. Non-Market Factors ................................................................. 51 

6. Social Factors .......................................................................... 52 

7. Cultural Factors ....................................................................... 52 

 

IV. Environmental Consequences ............................................................. 55 

A. Program Alternatives ...................................................................... 55 

1. Alternative 1: No Action ........................................................... 55 

2. Alternative 2: Removal of Infested Trees ................................ 63 

3. Alternative 3: Full Host Removal ............................................. 71 

4. Alternative 4: Infested Tree Removals and Insecticide 

 Treatment of High-Risk Host Trees .......................................... 75 

 5. Alternative 5: An Integrated Approach (Preferred 

Alternative) ................................................................................ 83 

 6. Cumulative Impacts .................................................................. 85 

B. Special Programmatic Considerations ........................................... 90 

1. Applicable Environmental Statutes ........................................... 90 

2. Mitigation and Risk Reduction .................................................. 98 

3. Program Monitoring .................................................................. 99 

4. Irreversible and Irretrievable Commitment of Resources ....... 100

 

Appendices 

A. Preparers 

B. List of ALB NEPA Links 

C. ALB Annotated Host List 

D. Assessment of Greenhouse Gas Fluxes and Climate Change Impacts 

Related to ALB 

E. Herbicide Human Health and Ecological Risk Assessment 

F. Imidacloprid Human Health and Ecological Risk Assessment 

G. Response to Comments on the Asian Longhorned Beetle Cooperative 

Eradication Program, Draft Environmental Impact Statement, March 

2015 

H. Acronyms and Glossary 

I. References 

J. Distribution List 

K. Index
Table of Contents iii.

Tables 

2–1 Overview of Alternatives in this EIS. 

3–1 GHG Emissions in the United States Allocated by Source. 

3–2 Threatened or Endangered Species that Occur in Areas the 

Program is Operating. 

3–3 Volume of Sawtimber (ALB-Host Species) in Northeastern States 

and the State of Ohio—2012. 

3–4 Volume of Roundwood Products (ALB-Host Species) in the 

 Northeastern States—2006. 

3–5 Value of Selected Timber Products for Species and State—2006 

3–6 Maple Syrup Production 2010–13. 

4–1. ALB Eradication Program Historical Cost-Share Contributions 

(from 1997 to 2009 (thousands of dollars)). 

 

Figures 

I–I Adult Asian longhorned beetle. 

I–2 Susceptibility potential for the Asian longhorned beetle in the 

United States based on the basal area for Acer (maple). 

3–1 Forest areas potentially at risk to Asian longhorned beetle 

3–2 U.S. hardwood log exports for select ALB-host species, 1989–2008. 

4–1 Direct and indirect effects of ALB establishment to representative 

aquatic and terrestrial organisms. 

4–2 Federally listed threatened and endangered species by county that 

co-occur with ALB-host trees in the Northeastern United States. 
iv. Table of Contents 

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Executive Summary v.

Executive Summary 

The U.S. Department of Agriculture, Animal and Plant Health Inspection 

Service (APHIS) has prepared a final programmatic environmental impact 

statement (EIS) for the Asian Longhorned Beetle (ALB) Eradication 

Program. ALB is a serious insect pest of certain hardwood tree species, 

with the potential to cause significant economic and environmental 

impacts if allowed to establish and spread throughout the United States. 

The Program is a cooperative effort between Federal and State agencies to 

identify and eradicate ALB infestations in the United States. To date, 

there have been ALB outbreaks in five States including Illinois, 

Massachusetts, New Jersey, New York, and Ohio. This final ALB EIS 

considers potential environmental impacts from each of the alternatives 

proposed for the APHIS ALB Eradication Program, should ALB be 

discovered elsewhere in the continental United States. APHIS can tier 

subsequent site-specific environmental assessments (EAs) to this EIS, 

incorporating, by reference, analyses included in this document, thus 

reducing response time for APHIS to act on new detections. In addition, 

this EIS will provide the interested public with a programmatic analysis of 

the potential for environmental impacts from the alternatives available to 

APHIS to eradicate ALB from the United States. 

On August 16, 2013, APHIS published a notice in the Federal Register 

(FR) describing its intent to prepare a programmatic EIS for the ALB 

Eradication Program (78 FR 50022–23, August 16, 2013). The public was 

invited to comment on the proposed EIS; APHIS received 27 comment 

letters during the 45-day scoping period. Comments were received from 

the public, non-governmental organizations, and State agencies regarding 

different aspects of the ALB Eradication Program. On March 16, 2015, 

APHIS published the draft EIS for public comment in the FR and 

notified interested parties of its availability. APHIS received 14 

comment letters during the 45-day comment period. APHIS has 

addressed the comments to the extent possible in this final EIS. 

Five alternatives were evaluated in this final EIS. The analysis is a 

general assessment of the alternatives, and their potential impacts to 

human health and the environment. The five alternatives include: 

(1) No Action—APHIS would implement quarantine restrictions but no 

eradication program; 

(2) Removal of Infested Trees—APHIS would implement quarantine 

restrictions and remove only ALB-infested trees; 
vi. Executive Summary 

(3) Full Host Removal—APHIS would implement quarantine restrictions 

and remove only ALB-infested trees, and all high-risk host trees up to a 

1⁄2 mile radius of infested trees; 

(4) Insecticide Treatment—APHIS would implement quarantine 

restrictions, remove all infested trees, and chemically treat all high-risk 

host trees with an insecticide up to a 1⁄2-mile radius of infested trees; 

and 

 

(5) An Integrated Approach—APHIS would implement quarantine 

restrictions, remove infested trees, and use a combination of removal 

and insecticide treatments of high-risk host trees (preferred 

alternative). 

The potential impacts from the implementation of the five alternatives 

suggest that there could be some effects to the human environment. The 

largest impacts, both economic and environmental, are expected to occur 

under the no action alternative, which would effectively allow ALB to 

become established and spread throughout the United States. Economic 

impacts related to the establishment of ALB in the United States in 

forested areas include a decrease in revenues related to various timber 

markets, maple syrup production, and tourism, where ALB-host trees are a 

primary component of impacted forests. Economic losses related to ALB 

establishment in urban areas include tree replacement costs and impacts to 

property value. Environmental impacts are anticipated for terrestrial and 

aquatic wildlife that are dependent upon ALB-host trees. The extent of 

the impacts will vary based on the prevalence of ALB-host trees in forests, 

and whether affected species depend on ALB-host trees to meet critical 

periods in their life history. The implementation of quarantine, in 

association with an eradication program, will reduce these impacts to 

varying degrees, depending on which eradication approach is used. 

The potential impacts from the proposed alternatives, and applicable 

environmental laws and statutes, are discussed on a programmatic basis 

in this EIS. No site-specific eradication projects will be implemented 

as a direct result of the decision that will follow this EIS. The decision 

to implement any treatment project will be made after site-specific EAs 

are conducted and documented, in accordance with the implementing 

procedures of the National Environmental Policy Act of 1969 (NEPA). 

Site-specific EAs will evaluate similar topics as there may be changes 

over time in the available data regarding this analysis, as well as 

applicable laws and statutes. 

Selection of the preferred alternative allows the Program to implement a 

proven eradication program that has been successful in other ALB 

eradication efforts in the United States. The preferred alternative 

integrates survey and quarantine with the removal of infested trees, and 
Executive Summary v.

site-specific management of high-risk host trees, thus allowing the 

Program the greatest flexibility in responding to ALB outbreaks and 

achieving eradication. 

I. Purpose of and Need for Action 1 

I. Purpose of and Need for Action 

The Asian longhorned beetle (ALB), Anoplophora glabripennis 

(Motchulsky), is a foreign wood-boring beetle that threatens a wide variety 

of hardwood trees in North America. The introduction of ALB into the 

United States was likely from infested wood pallets, or other wood 

packaging material (WPM), accompanying cargo shipments from Asia. 

The purpose of the proposed action is to protect the forests and trees of the 

United States from the adverse effects of ALB. 

Why is there a need to eradicate this pest? 

There is a need to eradicate ALB wherever it occurs because it is 

potentially one of the most destructive and costly invasive species to enter 

the United States. The beetle bores through the tissues that carry water 

and nutrients throughout the tree, causing the tree to weaken and 

eventually die. Symptoms occur approximately 3 to 4 years after 

infestation, and tree death can occur in 10 to 15 years, depending on site 

conditions. Infested host trees do not recover and regenerate (APHIS, 

2009). Tree mortality caused by ALB has been noted in countries where 

the beetle is endemic, and where it has been introduced (Haack et al., 

2010). In the United States, foresters have observed ALB-related tree 

mortality in New York, New Jersey, Massachusetts, Ohio, and Illinois. 

The insect threatens urban and suburban shade trees, and recreational and 

forest resources valued at hundreds of billions of dollars (Nowak et al., 

2001). In addition, ALB is likely to have negative impacts on forestdependent terrestrial and aquatic species, including threatened and 

endangered (T&E) species; soil and water quality could also be 

significantly impacted in forested areas where ALB-host trees are 

dominant (APHIS, 2009). 

Who has authority to act? 

The U.S. Department of Agriculture (USDA), Animal and Plant Health 

Inspection Service (APHIS) has a broad mission area that includes 

protecting and promoting U.S. agricultural health, and protecting and 

promoting food, agriculture, natural resources, and related issues. 

Specifically, the Plant Protection Act of 2000 (7 United States Code 

(U.S.C.) 7701 et seq.) provides the authority for APHIS to take actions to 

exclude, eradicate, and control plant pests, including ALB. Under this 

authority, APHIS works to prevent new infestations of ALB from entering 

the United States by regulating WPM, by restricting movement of items 

that may be infested with ALB (known as regulated articles) from areas 

under quarantine for ALB, and by conducting programs to eradicate ALB 

where it is found in the United States. 
2 I. Purpose of and Need for Action

This programmatic environmental impact statement (EIS) discloses the 

different methods and alternatives that APHIS could use to eradicate ALB 

from areas it occurs in the contiguous United States. 

Why do this environmental impact statement?

As a Federal Government agency subject to compliance with the National 

Environmental Policy Act of 1969 (NEPA) (42 U.S.C. 4321–4347), 

APHIS prepared this EIS in accordance with the applicable implementing 

and administrative regulations (40 Code of Federal Regulations (CFR) §§ 

1500–1508; 7 CFR §§1b, 2.22(a)(8), 2.80(a)(30), 372). This 

programmatic EIS presents program alternatives APHIS could adopt as 

part of the ALB Eradication Program, and examines the potential 

consequences of implementing them. 

APHIS has prepared 14 site-specific environmental assessments (EAs), 

since 1996, for ALB eradication programs and research studies in Illinois, 

New York, New Jersey, Massachusetts, and Ohio (appendix B). This EIS 

will consider potential environmental impacts from the APHIS ALB 

Eradication Program should ALB be discovered elsewhere in the 

contiguous United States. APHIS can tier subsequent site-specific EAs to 

this EIS, incorporating by reference analyses included in this document, 

thus reducing response time for APHIS to act on new detections, should 

these occur. In addition, this EIS will provide the interested public with a 

programmatic analysis of the potential for environmental impacts from the 

alternatives available to APHIS to eradicate ALB from the United States. 

A. Background 

a. Life Cycle 

ALB is in the wood-boring beetle family Cerambycidae. Adults are 1 to 

11⁄2 inches in length with long antennae, and are shiny black with small 

white markings on the body and antennae (figure 1–1). 

Adult females chew depressions into the bark of various hardwood tree 

species in which they lay (oviposit) their eggs. Eggs hatch within 2 weeks 

of oviposition (Haack et al., 1997; EPPO, 1999). After hatching, small 

white larvae feed within the vascular layer which transports nutrients and 

water throughout the tree and eventually bore into the tree. Larvae molt 

and can go through as many as 13 instar stages (Keena, 2008). The larvae 

continue to feed deeper into the tree's heartwood, forming tunnels (or 

galleries) in the trunk and branches. 

 

1. Description 

of ALB 
I. Purpose of and Need for Action 3 

 

 

 Figure 1–1. Adult Asian longhorned beetle. 

 

Over the course of a year, a larva will mature and then pupate inside the 

tunnels formed by the larva within the tree. The pupal stage lasts 13 to 

24 days (Lingafelter and Hoebeke, 2002; Haack et al., 2006). From the 

pupa, an adult beetle emerges, chewing its way out of the tree, forming 

characteristic round exit holes approximately 3⁄8 inch in diameter. After 

emerging from the tree, adults will feed on leaves and new growth bark for 

10 to 14 days before mating and laying eggs (Keena, 2002; Smith et al., 

2002). Adults tend to remain on the tree from which they emerge, 

resulting in infestation by future generations (EPPO, 1999). Adult activity 

is usually observed from May to July, although adults have been observed 

from April to December (Haack et al., 2010). 

The damage from larvae burrowing into the tree and adults burrowing out 

of the tree cuts off nutrient flow and weakens the tree. The tree will 

eventually die if the infestation is severe enough. Sawdust-like or thin 

wood shaving debris and insect waste (called frass) from the insect’s 

burrowing activity may be found at the base of infested trees. Infested 

trees are also prone to secondary attack by disease and other insects. 

In most locations, ALB produces one generation per year; however, in 

other countries where the pest is established, the number of annual 

generations varies with climate and latitude. For example, in northern 

China, one generation takes 2 years to develop (EPPO, 1999; Lingafelter 

and Hoebeke, 2002). Because ALB can overwinter in multiple life stages 

(egg, larval, and possibly pupal stages), adult emergence is staggered over 

time, resulting in adults emerging, feeding, mating, and laying eggs 

(ovipositing) throughout the summer and fall (Haack et al., 2006).

 

b. Hosts

 

There are 12 genera of host trees that APHIS regulates for ALB and are 

considered at-risk hosts: Acer (maple and box elder), Aesculus (horse 
4 I. Purpose of and Need for Action

chestnut and buckeye), Albizia (mimosa), Betula (birch), Cercidiphyllum 

(katsura tree), Fraxinus (ash), Koelreuteria (golden raintree), Platanus 

(sycamore and London planetree), Populus (poplar), Salix (willow), 

Sorbus (mountain ash), and Ulmus (elm) (7 CFR 301.51–2). These trees 

are hosts because the ALB can derive its food supply and complete its life 

cycle. A host tree is still a host even if it is not infested. Acer is the most 

commonly infested tree genus in the United States, followed by Ulmus and 

Salix (Haack et al., 2010). Acer, Aesculus, Betula, Salix, and Ulmus are 

considered preferred ALB-hosts (APHIS, 2008). Other regulated genera 

are considered rare or occasional hosts, including Albizia, Cercidiphyllum, 

Fraxinus, Koelreuteria, Plantanus, Populus, and Sorbus (APHIS, 2008) 

(see appendix C). A recent study indicates that ALB egg laying 

(oviposition) and reproductive success are higher on red maple (A. 

rubrum), compared to Norway maple (A. platanoides) or sugar maple (A. 

saccharum) (Dodds et al., 2014). In 2015, APHIS removed Celtis from 

the list of regulated host trees based on observation that this is not a host 

tree. (See appendix C for an annotated list of ALB hosts.) 

c. Dispersal and Spread 

(1) Human-Facilitated Spread 

ALB entered the United States, most likely, in WPM from Asia. USDA– 

APHIS regulations in 7 CFR §319.40–1 define WPM as “wood or wood 

products (excluding paper products) used in supporting, protecting, or 

carrying a commodity (includes dunnage).” WPM (e.g., pallets, crates, 

etc.) is made from low-grade lumber with higher moisture content than 

seasoned wood or heat-treated wood. Low-grade lumber may be of low 

quality due to pest damage (Cavey, 1998). Bush et al. (2002) reported that 

hardwood species accounted for about two-thirds of the total wood used 

for pallets during the 1990s. The National Plant Protection Organizations 

(NPPO) recognize WPM as a pathway for pest movement between 

countries. 

 

In response to the threat posed by untreated WPM, the Commission on 

Phytosanitary Measures of the International Plant Protection Convention 

(IPPC) of 2002 adopted International Standards for Phytosanitary 

Measures No. 15 (ISPM 15) titled “Guidelines for Regulating Wood 

Packaging Material in International Trade” (FAO, 2002 ). ISPM 15 

prescribes either heat treatment or fumigation for all WPM to mitigate risk 

of pests. WPM subjected to these approved measures is required to 

display a mark verifying compliance. Since September 16, 2005, the 

United States began full enforcement of ISPM 15, requiring either 

fumigation or heat treatment for all WPM entering the United States from 

any country; however, compliance with ISPM 15 is not required for WPM 

used domestically. WPM made from infested wood and used in domestic 

shipments would continue to be a potential source of spread. 
I. Purpose of and Need for Action 5 

Once ALB is present in the United States, it can move through domestic 

pathways. For example, movement of recently felled trees could carry 

ALB, and potentially spread it a long distance from the infested site. The 

level of risk depends on the end use of felled trees, that is, chipping for 

mulch, disposal in landfills, incineration, and so on. Although ALB 

prefers to infest branches in the tree crown rather than the trunk, lumber 

and saw logs are, nonetheless, a potentially important pathway to spread 

ALB. 

The production and distribution of fuel wood (this category includes 

firewood) also could transport infested material long distances. Moreover, 

many people use fallen limbs or felled trees from their private property as 

firewood, sometimes transporting and using these on vacations or in other 

homes. There is evidence that insect pests, including ALB, do spread 

through the movement of infested firewood (Koch et al., 2012). One 

estimate is that “3-5% of campground visits pose a potential risk of 

firewood-mediated dispersal of forest pests”, validating the concerns 

regarding the movement of pests through firewood (Koch et al., 2012). 

While educational programs and quarantine regulations in affected areas 

seek to prevent this kind of spread, this transport may still occur (Haack et 

al., 2010). While educational programs and quarantine regulations in 

affected areas seek to prevent this kind of spread, this transport may still 

occur. Transport of infested stumps and limbs for disposal are another 

potentially important ALB pathway, although the level of risk depends on 

how municipal solid waste is treated in the infested area. The movement 

of infested nursery stock can also transport ALB. 

Establishment of quarantines and quarantine regulations, which restrict the 

movement of woody material, reduce the likelihood of human-mediated 

pread of ALB. In quarantine areas in the United States, businesses or 

individuals working with trees or related products (e.g., firewood) must be 

under a compliance agreement with APHIS. These include landscapers, 

tree pruning and tree removal companies, firewood dealers, pallet 

distributors, nurseries, and sanitation workers, as well as other municipal 

or community services and associated contractors. The ALB Eradication 

Program provides a debris disposal site for use by companies under 

compliance agreement, and by municipalities within the quarantine area.

All host material for disposal, regardless of infestation, must be chipped or 

incinerated prior to leaving the quarantine area, or it must be permitted 

prior to being moved to other areas for treatment. Chipping infested trees 

is effective in destroying ALB (Wang et al., 2000). Nursery stock in the 

quarantine area is subject to inspection. Uninfested host material in the 

nursery trade may leave the quarantine area if accompanied by a 

certificate of inspection and approved permits. Any infested host material 

found in the nursery trade must be chipped. 
6 I. Purpose of and Need for Action

(2) Natural Spread

Adult beetles are prompted to move when the density of beetles in a given 

area reaches high levels (Williams et al., 2004), or when they are in search 

of a mate (Bancroft and Smith, 2005). ALB is a weak flier therefore 

spread through natural means (self-motility) is a minor contributor to 

spread potential (Bancroft and Smith, 2005). 

Various studies have examined the distance ALB travels per day. Smith 

et al. (2001) found, on average, that adult females and males move 

23 meters/day (m/day) and 17 m/day, respectively in a fragmented 

landscape composed dominantly (72%) of host trees planted in wind-rows 

bordering agricultural fields and dwellings in China. The researchers 

chose this site as it closely resembled the fragmentation in infested urban 

areas in the United States. Dispersal rates may be affected by the age of 

the beetle, population density, temperature, proximity of host trees, and 

other factors. In a mark-recapture study, Smith et al. (2004) found the 

median daily dispersal rate for both sexes was 30 m/day (range of 11 to 

56 m). Williams et al. (2004) tracked adult beetles using harmonic radar 

along a single straight row of 200 willow trees planted alongside a fourlane highway with agricultural fields on the other side. They found 

movement rates were 1.9 m/day for females and 3.7 m/day for males. 

Most beetles tended to move in one direction and rarely backtracked. 

Bancroft and Smith (2005) estimated a probability of 62% that beetles 

move from the tree on which they were last recorded to a tree nearby. 

They characterized daily movement at 20 m/day in a fragmented 

landscape in China with host trees lining roads, small fields, and houses as 

in areas of infestation in North America. 

Smith et al. (2001) measured a maximum ALB dispersal distance 

(including that of female beetles carrying mature eggs) of 1,442 m. In 

another mark-recapture study, Smith et al. predicted that “98% of beetles 

were recaptured at distances <920 m from the release point.” Beetles were 

recaptured at the most outlying sampling points (1,000 to 1,080 m from 

release point), and even at distances up to 2,600 m from the release point 

(Smith et al., 2004). Of the gravid (carrying eggs) females, 86 percent 

were recaptured within 1,080 m, and 77 percent were recaptured within 

600 m. Eight gravid females were recaptured at distances greater than 

2,000 m (2 kilometers (km)) (Smith et al., 2004). 

d. Geographic Distribution: Current and Projected

ALB is native to China and Korea (Cavey, 1998; Lingafelter and 

Hoebeke, 2002), but it was not considered a pest in natural forests there 

(Hajek, 2010). The distribution and abundance of ALB increased 

dramatically throughout China because in the 1960s and 1970s the 

Chinese government supported reforestation programs using fast-growing 
I. Purpose of and Need for Action 7 

poplar trees, many of which were susceptible to ALB (Cao et al., 2010; 

Hajek, 2010). Millions of ALB-infested trees have been cut in China over 

the past decades, and it is likely that this wood was used as WPM (Haack 

et al., 2010). Because of the increase in trade with Asia that involved 

WPM, ALB was introduced to Austria, Belgium, Canada, France, 

Germany, Italy, Japan, the Netherlands, Switzerland, the United Kingdom, 

and the United States (Carter et al., 2010; Haack et al., 2010; EPPO, 

2014). Belgium and Japan have eradicated the insect (EPPO, 2014). 

Peterson et al. (2004) predicted the potential distribution of ALB in North 

America based on the climatic factors in its native distribution. Their 

models suggest, “. . . the species has the potential to invade much of 

eastern North America, but only limited areas in western North America, 

and that a focus of initiation of invasion is likely to lie in the area south of 

the Great Lakes.” 

Haack et al. (1997) overlaid the distribution of ALB in China (21 °N to 

43 °N) onto North America, and determined that the range coincides with 

the span from southern Mexico to the Great Lakes. Keena (2006) noted 

that summer temperatures, throughout most of the lower 48 States, should 

support beetle survival and reproduction. Based on preferred host 

availability, Nowak et al. (2001) estimated the percentage of trees and 

number of trees at risk in nine U.S. cities, including one city in the 

Southeast and one city in the West. They estimated that 20 percent (or 

1.8 million) of trees in Atlanta, Georgia were at risk of infestation, and in 

Oakland, California, 12 percent (or 0.19 million) of trees were at risk. 

They also estimated that cities in the Northeast, such as Boston, 

Massachusetts and Syracuse, New York may lose as much as 60 percent 

of their trees should ALB be introduced and allowed to spread. All 

available studies indicate that ALB is able to survive and reproduce in any 

location within the lower 48 States that has host trees (APHIS, 2009). 

(See figure 1–2 for the ALB susceptibility potential for Acer (species of 

maple) in the United States.) 

The spread of ALB beyond the current isolated outbreaks could have 

severe economic effects. Although it is difficult to estimate the extent of 

expected impacts over time, it is clear that several industries in the 

Northeastern United States would be affected if ALB were to spread 

beyond the current areas of infestation to other urban and forested areas. 

These impacts would include not only the loss of urban trees at an 

estimated cumulative cost of at least $948 million in cities in the 

Northeastern United States, but also losses in several other industries, as 

well as non-market economic losses (Nowak et al., 2001). 

 

The timber industry produces over $178 million of products from host 

species annually in the Northeastern United States alone. Maple syrup 

production was worth nearly $67 million in 2010. The tourism industries 

2. Economic 

Impacts of 

ALB in the 

United States 
8 I. Purpose of and Need for Action

Figure 1–2. Susceptibility potential for the Asian longhorned beetle in the 

United States based on the basal area for Acer (maple). 

(Source: FS, 2009a). 

 in nine of the Northeastern States reported revenues around 

$143 billion in 2012. In addition, sales of deciduous shade trees were 

estimated at an additional $56 million for eight of the Northeastern States 

(including Ohio). The annual production values at risk in these three 

industries alone are significantly larger than the projected annual budget 

of the ALB Eradication Program. 

While the expected economic losses in forest-related products in the 

Northeastern United States and Ohio are large, previous research indicates 

non-market economic losses due to forest-related pests are expected to be 

larger than any market losses. This loss of host trees would have 

economic impacts in urban, suburban, and forested environments due to 

the loss of aesthetic and ecosystem benefits. Previous studies have shown 

that the presence of healthy trees increases property values by 2 to 6 

percent (Morales et al., 1976; Anderson and Cordell, 1988; Dombrow and 

Sirmans, 2000). In addition, the loss of host species would likely decrease 

fall foliage tourism in the Northeastern United States, which relies on a 

number of host species for the vibrant fall color displays. Many 

communities in the Northeastern States and the State of Ohio generate 

significant economic activity from fall foliage tourism, and would be 

unlikely to recover from the loss of so many of the species that provide fall 

colors. Other non-market impacts would be likely due to the loss of 

recreational benefits and ecosystem service provisions. The widespread 
I. Purpose of and Need for Action 9 

establishment of ALB would also likely affect U.S. trade in green lumber, 

and live trees and shrubs if importing countries establish quarantines or 

other restrictions for ALB. 

 

Although a complete cost-benefit analysis of the current Program is 

outside the scope of this EIS, even a relatively modest annual decrease in 

only two forest-related industries would result in economic losses more 

than sufficient to justify the anticipated costs of the ALB Eradication 

Program. The market losses in the timber and maple syrup industries do 

not include any impacts related to nursery stock or non-market losses 

related to urban trees and other aesthetic and ecosystem impacts. 

 

Considering that non-market losses are likely to be greater than any market 

losses, the benefits of an eradication program would outweigh program 

costs with a relatively small annual loss in the markets for forest products 

(APHIS, 2009). Further research to quantify these non-market benefits of 

eradication would help to strengthen this conclusion. 

 

Widespread establishment of ALB in the United States will cause 

significant ecological damage, affecting terrestrial and aquatic fauna and 

flora. Multiple factors dictate the extent of damage and its impact on 

ecological community function. In cases where the host trees are not a 

dominant component of the forest stand, the impacts may be less; 

however, in locations where host trees occur in high density and are 

keystone species within that particular forest type, the impacts are almost 

certain to be significant. The loss of keystone tree species and the 

associated cascading impacts to ecosystem function have been 

documented in other forest types where invasive forest insect pests were 

introduced (Ellison et al., 2005). The spatial and temporal factors related 

to an ALB infestation also will dictate the magnitude of impacts to 

ecosystems, particularly for species that are dependent on these habitats at 

critical periods in their life history. 

 

History of APHIS Eradication Actions and Other Regulatory 

Actions Against ALB

 

The first detection of ALB in the United States was in Brooklyn, New 

York, in 1996, although the actual date of introduction is much earlier. 

Shortly after, another infestation was detected in Amityville on Long 

Island. Since then, infestations have been found in Islip on Long Island, in 

Queens, in Manhattan and on Staten Island, including Pralls Island. Most 

recently, an infestation was detected in portions of Babylon Township on 

Long Island. 

 

Additional outbreaks occurred in Chicago, Illinois in 1998, and in 

Middlesex and Union Counties in New Jersey in 2002 (APHIS, 2007a). 

 

3. Existing and 

Projected 

Ecological 

Impact in the 

United States 

and 

Elsewhere 
10 I. Purpose of and Need for Action

In August 2008, ALB was found in Worcester County, Massachusetts and 

in July 2010, six ALB-infested trees were discovered in Suffolk County in 

Massachusetts. ALB was discovered in Clermont County, Ohio in June 

2011. 

 

In response to the introduction of ALB in these areas of the United States, 

APHIS implemented an adaptive management eradication program. 

Currently, the ALB eradication program in the United States uses— 

 

 visual surveys from the ground or aerial surveys by tree climbers or 

bucket trucks, 

 quarantine areas (up to 11⁄2 mile radius from infested trees) to restrict 

movement of regulated materials, 

 application of a preventative/protective systemic insecticide to uninfested trees, 

 removal of infested and high-risk host trees within the quarantine 

area, and 

 public involvement, outreach, and education about ALB. 

APHIS and its partners declared ALB eradicated from the following areas: 

 

 Chicago, Illinois and Hudson County, New Jersey, in 2008 

 Islip (Suffolk County, New York) in 2011 

 Manhattan (New York County, New York) and Staten Island 

(Richmond County, New York) in 2013 

 Middlesex and Union Counties, New Jersey in 2013 

 Boston (Suffolk County, Massachusetts) in 2014 

Active eradication programs are continuing in Queens, Brooklyn, Suffolk 

and Nassau Counties, New York as well as in Worcester County, 

Massachusetts, and Clermont County, Ohio. 

 
I. Purpose of and Need for Action 11 

B. Public Involvement 

APHIS has prepared 14 site-specific EAs regarding ALB eradication 

programs or research in New York, Illinois, New Jersey, Massachusetts, 

and Ohio, available for public comment since 1996 (appendix B).

APHIS provides many opportunities for public involvement and outreach 

regarding program activities in ALB-quarantined areas. As such, 

APHIS— 

 provided media interviews for newspapers, and radio and television 

outlets; 

 issued press releases; 

 conducted an annual advertising awareness campaign; 

 provided public service announcements on radio and television 

stations; 

 had a presence at industry shows, expos, and various outreach venues; 

 secured multiple airings on public television cable access stations 

across the United States of the “Lurking in the Trees,” ALB 

documentary, APHIS produced in conjunction with The Nature 

Conservancy, and made the documentary available on YouTube and 

iTunes; 

 posted information on social media including Facebook, Twitter, 

Pinterest and Flickr; 

 held public meetings as well as meetings with Federal and State 

officials, town administrators, and other impacted groups and persons; 

 provided informational materials and Web sites to the public, 

including an online reporting function and the arrangement of a 

national-use ALB hotline telephone number. 

Scoping is an open and early process to determine the issues to be 

addressed in an EIS, and to identify significant issues related to the 

proposed action covered in the EIS. As part of this process, APHIS sent 

out letters to all federally recognized tribal nations in the contiguous 

United States, inviting them to participate in a conference call on July 10, 

2013. During this conference call, APHIS asked tribal nations to provide 

input on how ALB and its eradication could affect them, and also provided 

opportunity for them to ask questions about the ALB Eradication Program. 

Approximately 20 tribes participated in the conference call. 

On August 16, 2013, APHIS published a notice of intent (NOI) in the FR 

describing its intent to prepare a programmatic EIS for the ALB 

Eradication Program (78 FR 50022–23, August 16, 2013). The public 

was invited to comment on the proposed EIS. 

1. Prior EAs 

Published by 

the Agency 

and Results 

from Public 

Comments 

3. Locations of 

Addressed 

Issues 

Raised by 

Stakeholders 

2. Past Agency 

Outreach 

Efforts 

on ALB 

Issues 
12 I. Purpose of and Need for Action

In the NOI, APHIS identified the following resources requiring further 

examination, in this EIS, of the potential environmental effects: 

 

 wildlife, including consideration of migratory bird species and changes 

in native wildlife habitat and populations; 

 federally listed T&E species; 

 soil, air, and water quality; 

 forests and trees in residential areas, and impacts on property values; 

 wood products industry and other economic impacts, including 

impacts on the firewood industry; 

 human health and safety; and 

 cultural and historic resources. 

APHIS made available a press release regarding the NOI to media 

contacts in New York, Massachusetts, and Ohio, and through the APHIS 

Stakeholder Registry that contains almost 12,000 contacts, and in the ALB 

e-newsletter on August 29, 2013. In addition, APHIS conducted the 

following notification activities: 

 ALB project managers in New York, Massachusetts, and Ohio 

shared the NOI with their staff members, and any key contacts that 

may have an interest in an ALB EIS; 

 Facebook post(s) on the ALB Facebook account located at 

https://www.facebook.com/asianlonghornbeetle; 

 Twitter post(s) on the ALB Twitter account located at 

https://twitter.com/StopALB;

 posted on the AsianLonghornedBeetle.com homepage and each State 

page; 

 posted on the APHIS ALB News and Information page; 

 Massachusetts State Department of Agricultural Resources and 

Department of Conservation and Recreation partners were notified and 

asked to share the NOI; 

 Ohio Department of Agriculture, Ohio Department of Natural 

Resources, Ohio State University, and Clermont County 

communication partners were notified and asked to share the NOI; 
I. Purpose of and Need for Action 13 

 notification to tribal contacts; 

 notification to USDA Forest Service (FS) contacts; 

 notification to U.S. Fish and Wildlife Service (FWS) contacts; 

 notification to various partners and organizations, such as the Tree 

Care Industry Association, Nature Conservancy, Arbor Day 

Foundation, American Forest Foundation, various arboretums, Sierra 

Club, Society of American Foresters, and so on; 

 APHIS–PPQ State Plant Health Directors in New York, Ohio, and 

Massachusetts shared the press release with FS, FWS, USDA– 

Agricultural Research Service (ARS), USDA–Natural Resource 

Conservation Service (NRCS), State agriculture/forestry/natural 

resource/heritage, and industry contacts. 

Although the comment period on the NOI was to end on September 16, 

2013, APHIS extended it to September 30, 2013. APHIS received 

27 comment letters during the 45-day scoping period. All comments were 

fully considered by APHIS in the planning of this EIS. Issues and 

concerns identified by the public and tribal contacts included: 

 

 impacts from heavy equipment use and other property damage 

from program activities; 

 impacts to soil and water quality; 

 impacts to maple syrup producers; 

 impacts to federally owned lands, such as national forests and national 

parks, State-owned nature preserves, natural heritage areas, and other 

high-value forest landscapes; 

 use of adaptive management as new studies or best management 

practices (BMPs) are identified; 

 insecticide pollution; 

 impacts to T&E species, migratory birds, and their habitats; 

 potential for ALB infestations through imported wood products; 

 biological control of ALB; 

 consideration of the Coastal Zone Management Act (CZMA) of 1972; 
14 I. Purpose of and Need for Action

 potential for yard waste and brush to spread ALB in the United States 

 risk of ALB to all of North America if ALB is allowed to spread; and 

 replacement of removed trees. 

Many comments did not raise specific issues for analysis in this EIS; 

however, opinions were provided for or against the selection of certain 

program control methods. 

APHIS and its cooperators recognize the public’s concern about the 

potential impacts of ALB and program activities on human health, 

biological resources, and the physical environment. Part of this EIS 

addresses these concerns. 

On March 16, 2015, APHIS published the draft EIS for public comment in 

the Federal Register (reference the FR) and notified interested parties of 

its availability. The comment period ended after 45 days. APHIS made 

available a press release regarding the draft EIS to media contacts in New 

York, Massachusetts, and Ohio, and through the APHIS Stakeholder 

Registry that contains over 12,000 contacts, and in the ALB e-newsletter 

on March 31, 2015. In addition, APHIS conducted the following 

notification activities: 

 APHIS-ALB project managers in Massachusetts, New York, and Ohio 

were asked to share the EIS with their staff members and any key 

contacts that may have an interest in an ALB EIS. 

 APHIS State Plant Health Directors (SPHD) in states with active ALB 

eradication programs were asked to share the draft EIS with their 

USDA-Forest Service (USFS), USDA-Agriculture Research Service 

(ARS), USDA-Natural Resource Conservation Service (NRCS), U.S. 

Fish and Wildlife Service (USFWS), State agriculture/forestry/natural 

resource/heritage, and industry contacts. 

 Information went to Media Update distribution list for Ohio ALB that 

was issued by the Ohio Department of Agriculture ODA. 

 Post on ALB Facebook account. 

 Post on ALB Twitter account. 

 Massachusetts Department of Agricultural Resources & Department of 

Conservation and Recreation partners notified and asked to share. 
I. Purpose of and Need for Action 15 

 Ohio Department of Agriculture, Ohio Department of Natural 

Resources (ODNR), Ohio State University, and Clermont County 

communication partners notified and asked to share. 

 New York Department of Agriculture and Markets, Department of 

Environmental Conservation and Parks partners notified and asked to 

share. 

 Email notification to Nature Conservancy contacts. 

 Email notification to USFS and USFWS contacts. 

 Email notification to various partners, such as TCIA, Arbor Day, 

American Forest, Arboretums, Sierra Club, Society of America 

Foresters, etc. 

 Notification to all State Plant Health Directors and State Plant 

Regulatory Officers (via GovDelivery as directed). 

 Notification to impacted members of Congress. 

 Notification to tribal contacts. 

 Ohio State University contact to send information to Ohio Horticulture 

List Serve, partnering organizations, associations and industry 

contacts. 

During the 45-day comment period, APHIS received 14 comment letters. 

Several commenters submitted their comments after the comment period 

closed. APHIS recognizes the importance of public comment and during 

the revision of the draft EIS, APHIS was able to review late comments as 

well. Appendix G provides a summary of APHIS’ response to these 

comments and how APHIS addressed the concerns in the final EIS. 

C. Decision Framework

 

In the earliest EAs for ALB eradications in Illinois and New York, 

alternatives available to APHIS included quarantine only, suppression, and 

eradication. Although eradication was the preferred alternative, no 

chemical methods of control were available at the time. Only host tree 

removal followed by burning or chipping of removed wood material was 

available. After 2000, APHIS added a prophylactic treatment using the 

insecticide imidacloprid as an option to the ALB Eradication Program 

after tests proved it effective in protecting trees from ALB under certain 

conditions. However, imidacloprid treatments do not ensure complete 

control of ALB within a tree due to variability in treatments, weather 
16 I. Purpose of and Need for Action

conditions, and tree health, all of which can result in uneven distribution 

of imidacloprid within a tree. 

In the most recent EA prepared for Clermont County, Ohio, in May 2013 

(APHIS, 2013b), APHIS analyzed four alternatives: (1) no action by 

APHIS; (2) removal of infested trees and high-risk host trees up to 1⁄2 mile 

from infested trees (full host removal); (3) removal of infested trees and 

imidacloprid treatment of high-risk host trees up to 1⁄2 mile from infested 

trees; and (4) infested host removal and combination of removal or 

imidacloprid treatment of high-risk hosts (preferred alternative). 

Alternatives 2 through 4 were eradication program options. 

Listed below are five alternatives for further examination in this EIS. 

(Chapter 2 describes the alternatives in greater detail.) 

1. No Action - Under the no action alternative, APHIS would not 

undertake eradication efforts. However, APHIS would continue to 

implement quarantine restrictions in the event of a confirmed ALB 

detection. This alternative represents the baseline against which a 

proposed action may be compared. 

2. Removal of Infested Trees - Under this alternative, APHIS would 

implement quarantine restrictions, and would only remove trees 

infested with ALB. High-risk host trees would not be removed or 

treated. 

3. Full Host Removal - Under this alternative, APHIS would implement 

quarantine restrictions, remove infested trees, and would remove 

high-risk host trees up to 1⁄2 mile from infested trees. 

4. Insecticide Treatment - Under this alternative, APHIS would 

implement quarantine restrictions, remove infested trees, and treat 

high-risk host trees with an insecticide up to 1⁄2 mile from infested 

trees. 

5. Integrated Approach (Preferred Alternative) - Under this alternative, 

APHIS would implement quarantine restrictions, would remove 

infested trees, and would use a combination of removal and 

insecticide treatments of high-risk host trees (preferred alternative). 

APHIS will not implement site-specific eradication projects as a direct 

result of the decision that will follow this EIS. Rather, APHIS will 

conduct site-specific EAs before the agency decides to implement any 

treatment project. EAs will address unique local issues, beyond the scope 

of this document, for site-specific management projects for ALB. Sitespecific EAs are more detailed and precise as to geographical locations 

and strategies appropriate for the type of outbreak. 
I. Purpose of and Need for Action 17 

The decision on this EIS will serve as the primary guide for management 

of ALB in the contiguous United States. Treatments and strategies allowed 

by prior EA decisions will continue to be available for use. The decision 

whether to plan or implement an ALB project in the United States will 

occur on a case-by-case basis by APHIS. 

D. Scope of this Document and NEPA Requirements

This EIS concerns only the APHIS ALB Eradication Program carried out 

by APHIS, directly or in conjunction with others (States, other Federal 

agencies, and tribal governments). The information and analysis 

contained in this EIS can be incorporated by reference into EAs and any 

other environmental documents prepared for ALB eradication projects, in 

accordance with NEPA. Some ALB-related activities, such as WPM 

importation regulations, inspection of WPM, and other ALB-regulated 

articles at the point of entry in the United States, and research and methods 

development activities are outside the scope of this document and were not 

examined. (For more information regarding environmental impacts of 

importation of WPM, see the APHIS EIS for Importation of WPM 

(APHIS, 2003), and the 2007 supplemental EIS (APHIS, 2007b)). 

E. Consultations 

Section 7 of the Endangered Species Act (ESA) and its implementing 

regulations require Federal agencies to ensure their actions are not likely 

to jeopardize the continued existence of T&E species, or result in the 

destruction or adverse modification of critical habitat. If necessary, 

APHIS conducts Section 7 consultation with the FWS and National 

Marine Fisheries Service (NMFS) on a site-specific basis for ALB 

eradication activities. 

APHIS considers whether critical habitat or listed species are present in 

the program area. If none are present, no Section 7 consultation is 

required. For the ALB Eradication Program in Worcester County, 

Massachusetts, APHIS consulted informally with FWS on a threatened 

plant, the small whorled pogonia (Isotria medeoloides), in 2008. In 2011, 

APHIS consulted informally with FWS on the impact of the ALB 

Eradication Program on the small whorled pogonia and piping plover 

(Charadrius melodus) in Norfolk and Suffolk Counties in Massachusetts. 

In June, 2011, APHIS first contacted FWS in Columbus, Ohio for 

technical assistance regarding impacts to federally listed species in 

Clermont County, Ohio. Seven endangered species (Indiana bat, Myotis 

sodalis; running buffalo clover, Trifolium stoloniferum; fanshell, 

Cyprogenia stegaria; rayed bean, Villosa fabalis; snuffbox, Epioblasma 

triquetra; pink mucket pearlymussel, Lampsilis abrupta; and sheepnose, 

Plethobasus cyphyus) occur in Clermont County. 
18 I. Purpose of and Need for Action

Since that time, APHIS submitted biological assessments and consulted 

with FWS on those species, receiving concurrence with “may affect, not 

likely to adversely affect” determinations, with the implementation of 

protection measures. In addition, FWS personnel have made site visits to 

the infested area, and have provided Indiana bat training to APHIS and the 

Ohio Department of Agriculture personnel. In addition, APHIS conducted 

surveys for Indiana bats in the Clermont County eradication work zones 

and reported those findings to FWS. Most recently, APHIS entered into a 

formal consultation with FWS on the Indiana bat (receiving a biological 

opinion dated June 4, 2014) and a conference on the northern long-eared 

bat (Myotis septentrionalis), a species proposed for listing as endangered 

(receiving a conference opinion dated July 3, 2014). The northern longeared bat was listed as threatened on May 4, 2015, and APHIS requested 

that FWS in Ohio convert the conference opinion into a biological 

opinion. In a letter dated May 1, 2015, FWS in Ohio confirmed the 

conference opinion for the northern long-eared bat as a formal 

consultation. In September 2014, APHIS requested a formal conference 

for the northern long-eared bat with FWS in New York for an ALB 

eradication on Long Island. When the northern long-eared bat was listed 

as threatened in 2015, APHIS then requested formal consultation for the 

bat. However, FWS indicated that informal consultation was more 

appropriate, and APHIS then submitted a request for informal 

consultation. APHIS received a concurrence letter from FWS in New 

York dated June 9, 2015. APHIS requested a formal conference with 

FWS in Massachusetts in November 2014. As listing of the bat 

approached, and discussions with FWS in Massachusetts took place, it 

was determined that an informal consultation was more appropriate. 

APHIS then submitted a request for informal consultation in April 2015, 

and received a concurrence letter from FWS in Massachusetts dated May 

14, 2015. 

APHIS will conduct ESA Section 7 consultations with the appropriate 

agency, as necessary, for any eradication programs if ALB is detected in 

new locations in the United States. Consultation with FWS, and NMFS, if 

necessary, at the local level will ensure that ALB eradication actions will 

not jeopardize the continued existence of a listed species or adversely 

modify critical habitat in the program area. APHIS will ensure the 

implementation of any protection measures for T&E species or critical 

habitat that result from such consultations. In addition, APHIS will ensure 

that site-specific consultations will be done, as necessary, under the 

National Historic Preservation Act, Migratory Bird Treaty Act, Bald and 

Golden Eagle Protection Act, and any other laws, regulations, Executive 

orders, and agency policies that apply to site-specific projects. 
II. Alternatives 19 

II. Alternatives 

This EIS analyzes the potential environmental consequences associated 

with the alternative options to eradicate ALB from areas in the contiguous 

United States where the insect establishes. The purpose of the alternatives 

is to describe the reasonable strategies the agency could take to achieve its 

goal. 

APHIS conducts survey activities and imposes Federal quarantines, 

according to the agency’s authority under the Plant Protection Act. 

Survey and quarantine are part of each alternative proposed, and are not 

unique to any one alternative. Therefore, descriptions for these two 

activities are independent from the descriptions for each alternative. 

A. Quarantine

Federal quarantine authority for ALB includes 7 CFR § 301.51 for 

eradication programs, 7 CFR § 319.40 for solid wood packing material, 

and 7 CFR § 330 for plant pests. Under these regulations, APHIS 

establishes quarantines and regulates international and interstate 

movement of regulated plant host material, also referred to as regulated 

articles. APHIS cannot regulate intrastate movement without the State 

Plant Regulatory Agency first establishing an intrastate quarantine. 

Intrastate quarantine facilitates regulatory activities within a geographical 

area less than an entire State. 

APHIS and State plant regulatory agencies establish quarantine boundaries 

1.5 miles from a tree with ALB-exit holes, and 0.5 miles from a tree with 

egg sites only. 

Under quarantine, APHIS restricts the movement of articles, known as 

regulated articles, which present a risk of spreading ALB interstate from 

the quarantine area. The regulated articles listed under the quarantine 

7 CFR § 301.51, as published in 2013, include the beetle and all its life 

stages; firewood (all hardwood species, not restricted to ALB-host trees); 

green lumber, and other living, dead, cut, or fallen material, including 

nursery stock, logs, stumps, roots, branches, and debris from ALB-host 

trees of 1⁄2 inch or more in diameter. 

As of April 2015, 12 genera of trees in the United States are regulated 

hosts for ALB, and are considered high-risk hosts because of ALB’s 

preference for feeding and completing its life cycle on these genera: Acer 

spp. (maple and box elder), Aesculus spp. (horse chestnut and buckeye), 

Albizia spp. (mimosa), Betula spp. (birch), Cercidiphyllum spp. (katsura 

tree), Fraxinus spp. (ash), Koelreuteria spp. (golden rain tree), Platanus 

spp. (sycamore and London planetree), Populus spp. (poplar), Salix spp. 

(willow), Sorbus spp. (mountain ash), and Ulmus spp. (elm). ALB can 
20 II. Alternatives

complete its life cycle in these materials, and transport of these materials 

could spread ALB to non-infested areas. In 2015, APHIS removed the 

genus Celtis from the list of regulated hosts based on evidence that ALB is 

unable to complete its life cycle on species within this genus. (Appendix C 

provides additional information and references on the list of regulated host 

trees of ALB.) 

Regulated articles originating from within the quarantine area may move 

outside the quarantine area if accompanied by a certificate or limited 

permit, unless the USDA moves the articles for scientific or research 

purposes. Issuance of a certificate or permit verifies that the host material 

underwent inspection by an APHIS-approved inspector, and is apparently 

free of ALB. At this time APHIS has no approved regulatory treatments, 

(treatments applied to ALB-regulated articles), that would allow for the 

interstate movement of regulated articles, except for heat treatment for 

firewood. The quarantine does not limit the natural spread of ALB. (A 

full description of the quarantine is found in 7 CFR 301.51.) 

Removal of an area under quarantine occurs upon declaration that the area 

is free of ALB. APHIS makes this declaration when the final survey finds 

no ALB infestations after control activities, secondary surveys, and a 

minimum of 4 years has passed since the last evidence of ALB in an area. 

The basis of the minimum 4-year survey interval is the beetle’s life cycle 

length (1 to 2 years), and the minimum length of time for the beetle’s 

population to grow to detectable levels should the beetle still be present. 

To declare eradication, a final round of negative survey is needed at some 

time after surveys conducted as part of delimitation and control measures 

have failed to find ALB in the area. The final time interval selected would 

depend on the size/area of the original infestation, the prevalence of 2-year 

ALB life cycles (if any) in the program area, and other programmatic and 

logistic factors. (Maps of ALB quarantine areas in the United States are 

available from the Plant Health link at 

http://www.aphis.usda.gov/wps/portal/aphis/home)

 

B. Survey 

 

ALB Program inspectors conduct surveys to (1) determine the scope of the 

infestation; (2) establish the quarantine area; (3) determine that ALB has 

not spread outside of the established quarantine area and, if it has, to 

expand the quarantine area; and (4) determine when to release an area 

from quarantine. The types of surveys conducted in an area depend on the 

scope of the infestation, and the circumstances surrounding the infestation. 

Below is a general description of surveys; however, there is flexibility in 

the survey protocols. For example, the number of times ALB Program 

inspectors survey an area depends on forest and host composition, the 

degree of infestation, and the time it takes between survey cycles. The 
II. Alternatives 21 

control strategy employed in an area (e.g., removal and/or treatment of all 

host trees in an area) will affect the number of survey cycles. As new 

biological information becomes available on ALB and survey strategies 

evolve, APHIS expects the survey protocols for the ALB eradication 

program to change. Site-specific EAs for the ALB eradication program 

will indicate changes to the following survey description. 

The discovery of ALB triggers a Level 1 survey, also known as core 

survey, to determine the scope of the infestation. Before the survey 

begins, the APHIS National Identification Service confirms the suspect 

specimen as ALB through structural and/or genetic identification. Once 

confirmed, Program inspectors survey neighboring host trees for 

oviposition (egg) sites and exit holes that may indicate an ALB infestation. 

During the Level 1 survey, ALB Program inspectors visually look for 

signs of infestation on every regulated host tree in a circular radius around 

the infestation, until they fail to find ALB within approximately a 1⁄2-mile 

radius of infested trees. Signs of infestation on host trees include exit 

holes, oviposition sites, frass, tunneling (formed as larvae and bore into 

the tree), and sap flow from damaged sites. Inspectors search for signs 

using binoculars from the ground, and may conduct aerial tree inspections 

through trained professionals using bucket trucks to peer into trees from 

above, and by trained tree climbers to search for signs of an infestation 

within tree canopies. Interest groups and organizations voluntarily assist 

inspectors by searching trees from the ground using binoculars. Use of 

tree climbers is the most effective method of detecting signs of ALB; 

however, this is also a slower and more costly method (Hu et al., 2009). 

Currently, no method of visual survey, including visual surveys for ALB, 

is completely effective in pest detection (Bulman et al., 1999). 

If inspectors find additional infestations, APHIS extends the 1⁄2-mile radius 

from the outermost find. APHIS bases the 1⁄2-mile radius on studies 

published in the scientific literature, as well as unpublished data on ALB’s 

natural spread potential (Smith et al., 2001; Smith et al., 2004; Williams 

et al., 2004; Bancroft and Smith, 2005). 

Survey areas may include residential, commercial, or public land; access 

to these properties is necessary for the Program. The Program removes 

trees infested with ALB. 

The Level 2 survey is a safeguard to ensure that ALB is not spreading 

beyond the 1⁄2-mile radius around infested host trees established during the 

Level 1 survey. During the Level 2 survey, also known as a buffer survey, 

inspectors survey host trees within a minimum of 1 mile beyond the 

survey boundary set during the Level 1 survey. This results in 

approximately a 11⁄2-mile radius from the point of infestation for trees with 

less than 100 exit holes. Level 2 Survey areas should be expanded to a 

1. Level 1 

Survey 

(Core Survey) 

2. Level 2 

Survey 

(Buffer)
22 II. Alternatives

minimum of 2.5 miles from areas that are or were centers of high 

populations, as denoted by the presence of a cluster of trees with many 

exit holes or one or more trees with >100 exit holes. This boundary 

expansion only applies to survey and not regulatory boundaries (the two 

can differ). APHIS bases the additional 2.5 mile boundary expansion on 

the high infestation or population level which indicates ALB has likely 

been in the area for some time as well as the beetle’s behavior to move 

away from its hatch tree when ALB populations are high (Williams et al., 

2004). 

During the Level 2 survey, inspectors focus on maple trees (Acer spp.) 

where they are present. Additional preferred hosts (Aesculus, Betula, 

Populus, Salix, and Ulmus) are surveyed when maples are not present or 

are within 100 yards of the last inspected preferred host tree. Sweet birch 

(Betula lenta) and Japanese maple (Acer palmatum) are not considered 

preferred hosts. The Program may adjust the survey radius upon 

discovery of additional infested trees. 

Similar to the Level 1 survey, inspectors complete the survey using 

binoculars from the ground, and may use bucket trucks and/or tree 

climbers to perform aerial surveys. 

Depending upon terrain or geographic layout (e.g., urban/suburban streets 

versus a forested area), inspectors use differing techniques to move 

quickly through an area to determine the extent of the infestation and 

determine the quarantine boundary. For example, in an urban setting, 

inspectors may skip to the next street or block; in wooded locations, 

inspectors may use a grid pattern. The intention of this survey is to 

quickly understand the general area of infestation, and determine the type 

of response needed in terms of resources and quarantine enactment. 

The ALB Program may conduct a Level 3 survey, also called a high-risk 

site survey. This survey extends beyond established survey boundaries to 

inspect high-risk sites. 

The Program conducts high-risk site surveys (Level 3) to identify and 

inspect high-risk locations where potentially infested host material may 

have been transported, stored, processed, or sold. Site inspections 

primarily focus on maples when they are present. High-risk sites include, 

but are not limited to: 

 Landscape and nursery businesses 

 Tree and lawn care companies 

 Firewood producers and transporters 

 State and local parks/forestry departments 

 Local utility and sanitation services 

 Parks and campsites 

 Landfills and disposal sites 

3. Level 3 

Survey 

(High-Risk 

Survey) 
II. Alternatives 23 

 Import facilities that receive or have received high-risk cargo from 

known source countries 

 Stone dealers or stone cutting facilities, such as 

monument/headstone companies, near infestations because of 

the use of wood packing material to ship and store the stone 

 

Secondary surveys are conducted within existing Level 1 and 2 boundaries 

after an infestation boundary has been established (through Core and 

Delimitation Surveys), and infested and high risk host trees have been 

removed or treated. Secondary inspections are crucial to detecting 

residual ALB infestations since they tend to develop slowly and become 

easily visible and detectable over time. Secondary surveys continue until 

a final cycle confirms there are no infested tress. Completion of a final 

survey is dependent on several factors, including but not limited to: 

 Size and scope of the original infestation 

 ALB lifecycle models 

 Programmatic and logistic requirements 

 Environmental conditions 

To declare eradication, a final round of negative survey is required after 

delimitation, control activities, and secondary surveys have been 

completed. Four years is the minimum amount of time between an initial 

detection and a finale survey cycle is completed. When there is more than 

a four year lapse prior to the final survey cycle, an interim survey cycle 

may be needed. 

C. Alternatives Evaluated in this EIS

APHIS is considering five alternative options, four of which are 

eradication program options. The five alternative options derive from 

scientific research published in peer-reviewed scientific journals, and 

experience in the eradication programs in Illinois, New Jersey, New York, 

Massachusetts, and Ohio. These alternatives are not unique to this EIS; 

other EAs prepared by APHIS include one or more of these alternatives 

(appendix B). Briefly, the five alternative options considered in this EIS 

include No Action, Removal of Infested Trees, Full Host Removal, 

Insecticide Treatment, and An Integrated Approach (see table 2–1) 

Adaptive Management 

This EIS proposes the use of specific chemical treatments as part of the 

ALB Eradication Program under the various alternatives with the 

exception of the no action alternative. Herbicides are proposed for use in 

all the alternatives with the exception of the no action alternative, while 

insecticide use is proposed under alternatives 4 and 5. The Program could 

4. Secondary 

Survey 
24 II. Alternatives

add other treatment(s) that may become available in the future to currently 

approved treatments for managing ALB, referred to as adaptive 

 

Table 2–1. Overview of Alternatives in this EIS. 

Methods Alternatives 

1– 

No Action 

2 –

Removal 

of 

Infested 

Trees 

3 – 

Full 

Host 

Removal 

4 – 

Insecticide 

Treatment 

5 – 

Integrated 

Approach

Remove and chip or 

incinerate infested trees X X X X 

Remove and chip/incinerate 

high-risk host trees within a 

1⁄2-mile radius of infested trees 

 X X 

Grind stumps of cut trees X X X X

Treat stumps and sprouts of 

cut trees with herbicide X X X X 

Allow some stumps of host 

trees to regenerate X X 

Site restoration X X X X

Imidicloprid treatment of highrisk host trees within a 1⁄2-mile 

radius of infested trees (only 

with landowner’s permission) 

 X X 

* Imposing Federal quarantines and conducting surveys to determine the extent of ALB infestation 

are part of all the alternatives.

 

management. A new treatment would be available for use upon the 

agencies’ finding that the treatment is registered by EPA for use on ALB, 

and poses no greater risks to human health and nontarget organisms than 

are disclosed in this EIS for the currently approved treatments. The 

protocol for making the necessary finding that a treatment is authorized by 

adaptive management is as follows: 

1. Conduct a human health and ecological risk assessment (HHERA). In 

this risk assessment, review scientific studies for toxicological and 

environmental fate information relevant to effects on human health and 

nontarget organisms. Use this information to estimate the risk to 

human health and nontarget organisms. Include these four elements in 

the HHERA: (a) hazard evaluation, (b) exposure assessment, (c) dose 

response assessment, and (d) risk characterization. The HHERA will 

do the following: 

• Identify potential use patterns, including formulation, application 

methods, application rate, and anticipated frequency of application. 

• Review hazards relevant to the human health risk assessment, 

including systemic and reproductive effects, skin and eye irritation, 
II. Alternatives 25 

dermal absorption, allergic hypersensitivity, carcinogenicity, 

neurotoxicity, immunotoxicity, and endocrine disruption. 

• Estimate exposure of workers applying the chemical. 

• Estimate exposure to members of the public. 

• Characterize environmental fate and transport, including drift, 

leaching to ground water, and runoff to surface streams and ponds. 

• Review available ecotoxicity data, including hazards to mammals, 

birds, reptiles, amphibians, fish, and aquatic invertebrates. 

• Estimate exposure of terrestrial and aquatic wildlife species. 

• Characterize risk to human health and wildlife. 

2. Conduct a risk comparison of the human health and ecological risks of 

a new treatment with the risks identified for the currently authorized 

treatments. This risk comparison will evaluate quantitative 

expressions of risk (such as hazard quotients), and qualitative 

expressions of risk that put the overall risk characterizations into 

perspective. Qualitative factors include scope, severity, and intensity 

of potential effects, as well as temporal relationships, such as 

reversibility and recovery. 

3. If the risks posed by a new treatment fall within the range of risks 

posed by the currently approved treatments, publish a notice in the FR 

of the agencies’ preliminary findings that the treatment meets the 

requirements of this alternative. The notice must provide a 30-day 

public review and comment period, and must advise the public that the 

HHERA and the risk comparison are available upon request. 

4. If consideration of public comment leads to the conclusion that the 

preliminary finding is correct, publish a notice in the FR that the 

treatment meets the requirements of this alternative and, therefore, is 

authorized by this alternative for use in the APHIS ALB management 

program. APHIS will make available to anyone, upon request, a copy 

of the comments received and the agencies’ responses. 

The decision to be made as a result of this EIS will be programmatic. 

Decisions to use specific treatments in projects (including new treatments 

authorized under the protocol in this alternative) will be made after sitespecific environmental analyses are conducted and documented, in 

accordance with the agency NEPA implementing procedures.

Under this alternative, activities for ALB would not involve APHIS. Other 

Federal or non-Federal entities may apply control measures; however, 

APHIS would not manage or fund these measures. Under this alternative, 

APHIS could conduct surveys to determine the extent of an infestation, 

and implement quarantine restrictions where ALB infestations occur. 

 

1. No Action 
26 II. Alternatives

This alternative represents the baseline against which to compare a 

proposed action. This alternative is not an eradication program. 

This alternative would consist of the following: 

 

 Implementing or maintaining the quarantine and 

expanding or contracting the quarantine area 

based on ALB’s distribution, as determined 

through survey 

 Removal of all infested trees 

 Chipping or incineration of cut trees 

 Grinding or herbicide treatment of stumps 

 Restoring the tree removal site 

All of the 

alternatives in this 

EIS include 

survey and 

quarantine 

 

After tree removal, the Program destroys cut trees through incineration or 

chipping to destroy the ALB life stages (larvae, pupae, and adults) that 

may be within those trees, thus eliminating potential adult beetle 

emergence, dispersal, and mating. Eggs may be able to survive the 

chipping process due to their small size; however, after hatching, larvae 

develop by feeding on the thin layer of generative tissue lying between the 

bark and the wood of a stem (cambium)—chipping makes this area 

unsuitable for the development of larvae (Wang et al., 2000). Chipping of 

trees to a size of less than 1 inch, in at least two dimensions, removes the 

risk of ALB (Wang et al., 2000) and the wood is no longer subject to 

Federal or State regulations. The Program chips trees in place or takes the 

tree to an approved establishment for chipping. 

ALB can reinfest stumps, above-ground roots, and any shoots that may 

grow from these tree parts, so it can be beneficial to kill stumps from cut 

trees. To kill stumps, APHIS may grind them or use herbicides depending 

on several factors. For example, if an area is inaccessible to equipment 

used for stump removal and grinding, or when the area is sensitive to 

erosion or compaction, herbicides would be used. The Program uses the 

herbicide triclopyr by spraying or painting the root collar area, the sides of 

the stump, and the outer portion of the cut surface, including the cambium, 

until thoroughly wet but not to runoff. Foliar applications of triclopyr 

mixed with two other herbicides, imazapyr and metsulfuron-methyl, are 

applied to sprouting foliage from stumps that remain after tree removal to 

prevent regrowth. APHIS would apply adaptive management as described 

earlier in this section of the EIS. In some locations, APHIS may leave 

stumps and allow regrowth, particularly in areas prone to soil erosion or 

sensitive wildlife habitats. 

 

After tree and stump removal from yards and landscaped settings, the 

Program restores the area through grading and planting groundcover 

2. Removal of 

Infested 

Trees 
II. Alternatives 27 

consistent with the area where the removals took place. This reduces the 

opportunity for invasive weeds to establish, and provides a groundcover 

that will help hold the soil in place. However, in woodlot settings a seed 

bank already exists in the soil that would result in rapid vegetation growth, 

therefore, planting groundcover may be unnecessary. 

APHIS will remove an area from quarantine as described previously. 

The Program, under this alternative, would consist of the following: 

 

 Implementing or maintaining the quarantine 

and expanding or contracting the quarantine 

area based on ALB’s distribution as 

determined through survey 

All of the alternatives 

in this EIS include 

survey and quarantine 

 Removing infested trees 

 Chipping or incineration of cut trees 

 Grinding or herbicide treatment of stumps 

 Restoring the tree-removal site 

As described in 

Alternative 2— 

Removal of Infested 

Trees 

 Removing high-risk host trees within a 1⁄2-mile 

radius of an infested tree(s) 

 

Under this alternative, the Program would conduct pest surveys, 

implement quarantine restrictions, and remove infested host trees, as 

described in alternative 2. Signs of low infestation levels may not be 

readily apparent on high-risk host trees, and can remain unnoticed by 

visual survey. Consequently, due to their proximity to known infested 

trees, there are chances that nearby high-risk host trees have undetected 

infestations or are at risk of infestation. Therefore, the Program would 

also remove high-risk host trees within a 1⁄2-mile radius of infested host 

trees. 

The basis for removing trees within the 1⁄2-mile radius is on the dispersal 

behavior of ALB, and the level of effectiveness of visual survey on lightly 

infested trees. In a study in Illinois, 99 percent of trees with ALB egglaying sites were within 1⁄4 mile of a tree from which adult ALB exited 

(Sawyer, 2006). APHIS may add an additional 1⁄4 mile to the radius for 

host tree removal around infested trees to capture beetles that may have 

spread further (Sawyer, 2006). APHIS will notify the landowner prior to 

the removal of infested and high-risk host trees. If the landowner refuses 

to allow removal of high-risk host trees, the Program would not remove 

them, but would continue to survey and, if ALB infests those trees, APHIS 

would remove the trees after notifying the landowner. 

The cutting and removal of trees is the same as described in alterative 2. 

The Program would focus tree removal efforts first by removing infested 

trees to eliminate the population of ALB, and then remove high-risk host 

trees located within the 1⁄2-mile radius of infested trees. However, tree 

3. Full Host 

Removal 
28 II. Alternatives

removal may not always occur in this order. For example, it is more time 

and cost effective to finish removing all the infested and high-risk trees in 

an area before relocating equipment to a new staging area. 

APHIS removes and grinds or treats stumps and roots of felled trees with 

herbicides, as described in alternative 2. However, in some cases, such as 

woodlots, stumps may not be ground or treated with herbicides to allow 

for regrowth of host trees. APHIS would apply adaptive management, as 

described earlier in this section of the EIS. 

APHIS will remove an area from quarantine as described previously. 

 

The Program, under this alternative, would consist of the following: 

 

 Implementing or maintaining the quarantine 

and expanding or contracting the quarantine 

area based on ALB’s distribution, as 

determined through survey 

All of the 

alternatives in this 

EIS include survey 

and quarantine 

 Removing infested trees 

 Chipping or incineration of cut trees 

 Grinding or herbicide treatment of stumps 

 Restoring the tree-removal site 

As described in 

Alternative 2—

Removal of Infested 

Trees 

 Treating high-risk host trees within a 1⁄2-mile 

radius of infested trees with an approved 

insecticide, following product label 

requirements 

 

Under this alternative, APHIS would remove only infested trees, as 

described in Alternative 2, Removal of Infested Trees. Due to their 

proximity to known infested trees, there is a chance nearby host trees are 

at risk of infestation. APHIS would treat high-risk host trees, located up 

to 1⁄2 mile of an infested tree, with the insecticide imidacloprid, but only 

with permission from the landowner. APHIS uses the insecticide to 

protect trees from ALB infestation but does not use the insecticide to treat 

infested trees. If the landowner does not allow imidacloprid treatment of 

trees on their property, the Program will not treat them; however, survey 

efforts would continue. Once trees become infested with ALB, the 

Program would remove them following notification of the landowner. 

Application of imidacloprid is through trunk or soil injection, according to 

product label requirements. The Program makes insecticide treatments in 

the spring, early summer or fall, prior to and during the adult emergence 

period, in order to allow the insecticide to distribute throughout the tree 

and, therefore, be most effective. 

The rate of imidacloprid used depends on the application method, as well 

4. Insecticide 

Treatment 
II. Alternatives 29 

as the size of the tree to be treated. Pesticide applicators adhere to the 

requirements on the product label. For trunk injections, applicators drill 

holes around the trunk, 2 to 6 inches above the soil-wood line, and inject 

imidacloprid. It takes 1 to 3 weeks for the insecticide to distribute 

throughout the tree, depending on the size and condition of the tree and 

weather conditions. Trunk injections are the most common application 

used by the Program. For soil injection, applicators inject imidacloprid at 

a minimum of four injection sites, spaced evenly around the base of the 

tree. Application occurs under the soil around the base of the tree, 

normally no more than 12 inches from the base. No material may puddle 

or run offsite. It may take up to 3 months before plant tissues absorb 

sufficient quantities of imidacloprid, depending on the size and condition 

of the tree and weather conditions. 

For maximum efficacy, APHIS would ideally repeat insecticide 

applications once a year over a consecutive 3-year period to ensure that 

the concentration of the insecticide within the treated tree is at an adequate 

level to protect the tree from ALB infestation. Imidacloprid treatments do 

not ensure complete control of ALB within a tree due to variability in 

treatments, weather conditions, and tree health, all of which can result in 

uneven distribution of imidacloprid within a tree. APHIS would apply 

adaptive management as described earlier in this EIS. 

APHIS will remove an area from quarantine as described previously. 

An integrated approach would consist of the following: 

 Implementing or maintaining the quarantine 

restrictions, and expanding or contracting the 

quarantine area based on ALB’s distribution 

as determined through pest surveys 

All of the alternatives 

in this EIS include 

survey and quarantine 

 

• Removing infested trees 

• Chipping or incineration of cut trees 

• Grinding or herbicide treatment of stumps 

• Restoring the tree-removal site 

 Removing high-risk host trees within a 1⁄2-

mile radius of an infested tree(s)

 

 

As described in 

Alternative 2—

Removal of Infested 

Trees 

As described in 

Alternative 3—Full 

Host Removal 

• Removing or treating with an approved 

insecticide, according to label requirements, 

high-risk host trees up to 1⁄2 mile-radius from 

infested tree 

 

As described in 

Alternative 4—

Insecticide Treatment 

 

For this alternative, the Program will remove infested trees and treat with 

imidacloprid, or remove high-risk host trees within a 1⁄2-mile radius of 

infested trees. APHIS will base the determination to treat or remove high5. An Integrated 

Approach 

(Preferred 

Alternative) 
30 II. Alternatives

risk host trees on levels of infestation in the area, host tree density and 

distribution, potential environmental impacts, and logistical resources. 

This alternative provides the most flexibility in selecting an appropriate 

control method for a location. It is also the most cost effective method 

because this alternative does not prescribe that all high-risk host trees must 

be treated or removed; rather, it allows flexibility in focusing treatments 

on the host trees most preferred by ALB (i.e., Acer, etc.), or certain 

locations that would be higher risk than others. 

As with the other action alternatives, APHIS would remove or treat with 

imidacloprid the high-risk host trees located up to 1⁄2 mile of an infested 

tree only with permission from the landowner. If the landowner does not 

allow removal or imidacloprid treatment of trees on their property, the 

Program will not remove or treat them but will continue survey efforts. If 

trees become infested with ALB, the Program would remove them 

following notification of the landowner. 

APHIS would apply adaptive management as described earlier in this 

section of the EIS. 

APHIS will remove an area from quarantine as described previously. 

D. Alternatives Considered but Not Included in 

this EIS

APHIS evaluated other insecticide chemistries for their efficacy towards 

ALB (appendix B). Imidacloprid is currently the only effective, registered 

product available for use against ALB. 

a. Biological Control

Biological control (biocontrol) is the means of reducing or mitigating pests 

and pest effects with natural enemies. ALB causes damage within its 

native range despite the presence of associated natural enemies. Research 

on biocontrol methods for ALB are ongoing, however, none are available 

for large-scale use. 

b. Sterile Insect Technique

Sterile insect technique (SIT) involves breeding and releasing large 

populations of sterile male insects with the goal of outcompeting fertile 

males during mating with females. This leads to a reduction in the size of 

the population, and potential eradication of an insect population. Current 

SIT technology is not feasible based on the ALB life cycle. 

c. Insect Pheromones

1. Use of Other 

Chemical 

Control 

Agents 

2. Other 

Integrated 

Pest 

Management/ 

Suppression 

Strategies 
II. Alternatives 31 

Insect pheromones are naturally produced chemicals that can be used by 

insects to attract each other, and have been used in pest management for 

survey and suppression of target pests. The identification of insect 

pheromones specific to ALB is currently an ongoing area of research; 

however, additional work is needed before it can be successfully 

implemented in the ALB Eradication Program. 

d. Regulatory Treatments

APHIS has determined that it is necessary to add a treatment schedule for 

ALB in the Plant Protection and Quarantine Treatment Manual. A 

treatment evaluation document has been prepared that discusses the 

existing treatment schedule and explains why this change is necessary. 

The document was made available to the public for review and comment 

until April 11, 2014. The public can access the “Notice of Availability of 

a Treatment Evaluation Document for Heat Treatment for Asian 

Longhorned Beetle” (Docket No. APHIS-2013-0094) online at 

http://www.regulations.gov/#!docketDetail;D=APHIS-2013-0094). 

e. Additional Methods Development and Research

Additional methods development projects include looking at chip size and 

grinding techniques to deregulate host material, and assessing the use of 

pesticide treatments in the fall. Research continues with analysis to 

determine how fast the insect spreads on its own, evaluating the host trees 

the beetle attacks for preference and range, DNA analysis and behavioral 

experiments. 

 
32 II. Alternatives

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III. Affected Environment 33

III. Affected Environment

This section presents a baseline description of the environmental resources 

that would be affected by implementation of the alternatives presented in 

chapter 2. Environmental resources include physical and biological 

resources, as well as the economic and social factors affecting people. 

Because this is a programmatic EIS, the description of the affected 

environment contained in this chapter is, by necessity, general. The 

potentially affected environment in the United States is anywhere host 

trees susceptible for ALB infestation are found. Given the known 

distribution of ALB, the insect is probably capable of surviving anywhere 

in the United States where suitable host plants and climatic conditions are 

available. 

A. Affected Forest Communities

Forest trees grow either in pure stands comprised of a single species or in 

mixed stands. Tree density and type influence the sub-canopy plant 

species by altering the shade levels, soil composition, moisture levels, and 

other attributes. Forests fall on public or private lands, and occur on 

uplands, riparian areas (the interface between terrestrial and aquatic 

ecosystems), wetlands, and coastal environments. 

Indicators of forest conditions include tree mortality and growth rates, 

degree of insect and pathogen damage, and species composition in the 

understory and canopy. Tree mortality from ALB can affect the 

composition of forest communities. As ALB and other introduced insects 

and pathogens spread, they add stress to forest communities. Other forest 

stressors, including reduced water and air quality, and geological and soil 

disturbance often result from human activity. As the proximity and size of 

human populations increase, the forest canopy coverage and biological 

diversity decreases, while fragmentation and ecological disturbance 

increases. 

In the United States, 12 tree genera are regulated hosts for ALB: Acer 

(maple and box elder), Aesculus (horse chestnut and buckeye), Albizia 

(mimosa), Betula (birch), Cercidiphyllum (katsura tree), Fraxinus (ash), 

Koelreuteria (golden raintree), Platanus (sycamore and London 

planetree), Populus (poplar), Salix (willow), Sorbus (mountain ash), and 

Ulmus (elm) (7 CFR 301.51–2). The degree of susceptibility of a forest to 

ALB is dependent upon the composition of tree species, including the 

total density (basal area per acre) and proportion of an area covered by 

susceptible stands. Within the United States, the distribution of host 

species is largely concentrated in the mid-Atlantic, New England, and 

Great Lakes regions (referred to as the Northeast Region by the Forest 
34 III. Affected Environment 

Service (figure 3–1), occupying a range of ecosystems including riparian 

and wetland areas. In the Northeast Region, forests (mostly privately 

owned) cover 42 percent of the land base (Shifley et al., 2012). The 

composition of trees in the Northeast Region includes the maple-beechbirch forest-type group that accounts for 29 percent of the forest area; 

elm-ash-cottonwood for 6 percent; and aspen-birch for 10 percent (Shifley 

et al., 2012). Combined, these forest-type groups account for 45 percent of 

the forest area. In other regions of the United States, ALB-host species 

are found but make up a smaller component of the tree species growing in 

these forest areas (figure 3–1). However, many of these species are 

planted as ornamentals and urban shade trees elsewhere in the United 

States (Peterson et al., 2004). 

 

Figure 3–1. Forest areas potentially at risk to Asian longhorned beetle. 

**This map does not include host genera planted in urban 

and suburban areas. (Data source: USGS, 2014) 

 

It is possible that other tree species could become confirmed hosts for 

ALB, which may increase the geographic scope of the environmental 

impacts. (See appendix C for an annotated list of ALB hosts.) 

B. Affected Water Resources

Surface and ground waters within or near affected forest communities may 

be part of the affected environment. Surface waters include streams, 

lakes, wetlands, and reservoirs. Surface water seeps underground forming 
III. Affected Environment 35

aquifers, also known as ground water. The combination of surface 

waterand underlying ground water within an area that drains to a common 

outlet, (e.g., a waterway, wetland, reservoir, aquifer, or ocean) is a 

watershed. In the contiguous United States, there are 2,110 watersheds 

that may contain ALB-host trees. Nationally, around 53 percent of the 

water supply originates on forestlands (Shifley et al., 2012). In the 

Northeast Region, 48 percent of the water supply originates on 

forestlands; in the South Region, 51 percent; in the Pacific Coast Region, 

72 percent; and the Interior West Region, 43 percent (Shifley et al., 2012). 

American Heritage Rivers are rivers that represent natural, cultural, and 

historical resources and flow through forest ecosystems, including 

suburban and urban forests. Created by Executive Order 13061 

(September, 11, 1997), with selection criteria developed by the Council on 

Environmental Quality (CEQ), the EPA has designated 14 rivers in the 

contiguous United States as American Heritage Rivers. Examples include 

the Connecticut River, Cuyahoga River, Hudson River, Mississippi River, 

and the Potomac River. 

The National Estuary Program is a network of voluntary community-based 

programs which safeguards the health of important coastal ecosystems 

across the country. The National Estuary Program, established under 

Section 320 of the 1987 Clean Water Act (CWA) amendments, calls for 

each national estuary program to develop and implement a comprehensive 

conservation and management plan that contains specific targeted actions 

designed to address water quality and habitat challenges in its estuarine 

watershed (EPA, 2013c). Twenty-seven national estuary programs are in 

place throughout the contiguous United States.

The Coastal Zone Management Act of 1972 (as amended), administered 

by the U.S. Department of Commerce, National Oceanic and Atmospheric 

Administration (NOAA), Office of Ocean and Coastal Resource 

Management (OCRM), provides for management of the nation's coastal 

resources, including the Great Lakes. Federal agencies are to cooperate 

with NOAA if their activities affect any land or water use, or natural 

resource of the coastal zone (16 U.S.C. § 1456, Section 307). Host trees 

of ALB occur in some coastal zones; ALB infestation and pursuant 

management activities can affect natural resources in these areas. 

Physical, biological, and chemical alterations to ecosystems through 

natural means or invasive species and human activity can affect water 

quality and quantity. Indicators of water quality include the flow rate, as 

well as water and sediment chemistry. 

Forested landscapes directly and indirectly affect water resources through 

a variety of mechanisms, including the stabilization of soil and prevention 

of sediment runoff; influence on the water cycle by controlling rainfall 

1. Water 

Quality 
36 III. Affected Environment 

runoff, flooding, and uptake and evapotranspiration of water; and effects 

on water temperature and understory plant growth by casting shade. 

Canopy gaps (reduced shade cover) can contribute to a decrease in soil 

moisture due to the increase in levels of light reaching the ground, as seen 

with the loss of hemlock trees to the hemlock wooly adelgid, an insect pest 

introduced in the United States (Orwig and Foster, 1998; Orwig et al., 

2008). Canopy gaps can also contribute to an increase in water 

temperature thereby affecting aquatic life, including plants, animals, and 

invertebrates (Kaushal et al., 2010). Reduced tree density and canopy 

gaps can affect forest undergrowth and soil stability, and can increase 

overland water flow, water yield, and runoff; however, the magnitude, 

timing, and duration of the response varies considerably among forest 

communities (Hornbeck et al., 1993; Wear and Greis, 2013). Increases in 

water yield coincide with the mobilization and leaching of nutrients in 

some forest communities (Hornbeck et al., 1993). Urban development and 

conversion of forested lands to agriculture are major contributors to 

reduced forest canopies and increases in water runoff, affecting the 

temperature and water chemistry of streams and water bodies (Wear and 

Greis, 2013). 

The vegetation in riparian areas and wetlands contribute to water quality 

and quantity. A wetland is a land area saturated with water at a frequency 

and duration to support vegetation adapted for life in saturated soil 

conditions. Many of the tree species that are regulated hosts to ALB grow 

in riparian and wetland areas, including maple (Acer), elm (Ulmus), ash 

(Fraxinus), cottonwood (Populus), sycamore (Platanus), willow (Salix), 

and birch (Betula) (USGS, 1998; NHDES, 2008; NCFS, n.d.). Changes in 

tree cover and density in these aquatic habitats can have negative impacts 

on streambank stabilization, water temperature, sediment loading, 

hydrology, nutrient cycling, and contaminant removal (Wenger, 1999; Lee 

et al., 2004; Jones et al., 2006).

 

For example, maple trees are a critical component in soil nitrification in 

the Northeast United States, and their loss could affect nitrogen retention 

and cycling in forested watersheds (Lovett and Rueth, 1999; Lovett and 

Mitchell, 2004). Nitrification is the oxidation of nitrogen-containing 

compounds into nitrate, which is an important component of the nitrogen 

cycle. A decrease in maple stands would lead to higher nitrogen retention 

in soils and reductions in nitrogen transport into aquatic systems. The 

alteration of nitrogen cycling, due to the loss of maples, would alter plant 

succession and diversity in terrestrial environments, as well as affect 

aquatic ecosystems that are dependent on higher nitrogen inputs. 

Forested landscapes, including floodplains, wetlands, and riparian areas 

moderate flooding and filter sediments and pollutants (Dosskey et al., 

2010), protecting water quality. Alterations to the quality of these surface 

waters can impair the values they impart. Human communities rely on 
III. Affected Environment 37

surface and ground water for drinking water, irrigation, and recreational 

activities. Waters in urban and agricultural areas tend to have an increased 

nutrient concentration and load. Excess nutrients mostly arise from 

fertilizers, wastewater effluent, and industrial waste in urban areas; and 

animal waste, fertilizers, and chemicals in agricultural areas (Wear and 

Greis, 2013). 

The Clean Water Act (CWA) provides a structure for regulating the 

discharge of pollutants into waters, and regulates quality standards for 

surface waters. Section 303(d) of the CWA requires States to develop lists 

of its impaired and threatened waters (stream/river segments, lakes) (EPA, 

2014c). Causes for impairment are numerous and include categories such 

as pathogens, metals, salinity, sediments, pesticides, trash, and other 

organic and inorganic compounds. Most States have one or more 

impaired waters, meaning that the water is not meeting one or more of its 

designated uses (EPA, 2014c). Assessments are not complete for all 

watersheds. 

Soil quality impacts water quality and availability. A well-managed soil 

will have good porosity (space between soil particles), allowing it to be an 

efficient receiver of rainwater. Water that infiltrates the soil, in the 

absence of excessive nutrient or contaminant loads, is generally purified 

before entering ground water sources or returning to surface water bodies 

(Karlen et al., 1997). However, improperly managed or disturbed soil 

typically results in poor porosity, leading to surface water runoff carrying 

potential pollutants and soil particles with it. Soil enters surface waters as 

sediment, and can negatively affect water quality. Smaller particles (e.g., 

clay) stay in suspension contributing to water turbidity (Cook, 1990). 

Riparian and floodplain habitats are especially sensitive to changes in 

water quality (Doupé et al., 2010). By volume, sediment is the largest 

cause of impairment of rivers and streams across the United States 

(Cunningham et al., 2001; EPA, 2013b). 

C. Affected Soil 

Soil types capable of supporting ALB-host trees are part of the affected 

environment. The physical, chemical, and biological characteristics of soil 

affect the health of the vegetation it supports by changing the availability 

of water and nutrients. 

Soil is composed of a diversity of mineral and organic components. Soil 

stores, cycles, and moderates the release of nutrients and other elements. 

Soil sequesters carbon, reducing the prevalence of atmospheric carbon 

dioxide, a gas linked to climate change. In the Northeast, roughly as much 

carbon is stored in the soil, mostly in the form of organic matter, as in live 

biomass (e.g., trees) (Shifley et al., 2012). 

2. Interactions 

of Soils with 

Water Quality 
38 III. Affected Environment 

In most ecosystems, soil biota help to regulate a number of key ecosystem 

services, including plant production, nutrient and carbon cycling, 

maintenance of soil structure, and water regulation (Wall et al., 2012). Soil 

biota can have direct and indirect impacts on land productivity and other 

ecosystem functions, such as fresh water; food and pollination services; 

timber, fiber, and fuel; nutrient and waste management; and climate 

regulation (Barrios, 2007). 

Vegetation alters soil nutrient cycling, especially in situations where 

certain vegetation types serve a unique ecological function. For example, 

maple trees are a critical component in soil nitrification in the Northeast 

United States, and their loss could affect nitrogen retention and cycling in 

forested watersheds (Lovett and Rueth, 1999). 

Soils are categorized by type, which describe the physical properties of the 

soil (including permeability, water-holding capacity, soil texture, and soil 

structure), and its chemical properties (including pH, salinity, cation 

exchange capacity, organic matter, and carbon-to-nitrogen ratio). The 

Natural Resources Conservation Service categorizes soil type into 12 soil 

orders and 64 suborders (McDaniel, 2006). Based on their various 

characteristics, different soil orders and suborders have varying capacities 

to support ecosystem services (i.e., retain water, filter water impurities, 

cycle nutrients, anchor plant roots, and absorb air pollutants). As a result, 

disturbance affects different soil types to varying degrees and in a variety 

of ways. 

Changes in physical soil characteristics occur when ground-based 

equipment makes repeated passes over the soil. These activities compact 

soils and, if soils are wet enough, can result in rutting and leaching of soil 

nutrients. Different soil types compact more readily than others; clay and 

loam soils generally compact more than sandy soils. These changes to the 

physical soil characteristics reduce the pore space and the ability of the soil 

to retain water. In turn, this reduces infiltration rates, slows soil drainage, 

impedes root growth, and reduces plant-available water and nutrients. 

Physical soil disturbances also decrease gas exchange, affecting both 

plants and soil biota. 

Organic matter affects the chemical soil characteristics. In its various 

forms, organic matter provides nutrients and retains moisture for soil 

organisms and plants. Because organic matter derives from decomposing 

plant material, the type of vegetation growing in an area influences the 

chemical composition of soil. 

Soil quality refers to “the capacity of a soil to sustain biological 

productivity, maintain environmental quality, and promote plant and 

animal health” (Doran and Parkin, 1994). Erosion, compaction, loss of 

soil structure, loss of nutrient content, and changes in soil salinity degrade 

1. Soil Types 

of the 

United States

2. ALB-Host 

Species and

Soil Quality 
III. Affected Environment 39

soil quality (Cook, 1990). Atmospheric acid deposition has led to a 

concentration of highly acidic soils in the Northeastern States and 

southward along the Appalachian Mountains (Shifley et al., 2012). A 

continued increase in acidity is likely to make sensitive trees more prone 

to other stressors, for example, insect attacks. Soil types vary in response 

to different impacts and the preparation of EAs at the local or regional 

level for specific Program actions will address the site-specific issues. 

The types of trees that grow in a forest ecosystem affect the quality of the 

forest soil. Predominant or “foundation” species often define the structure 

of a forest community by creating locally stable conditions for other 

species, and by modulating and stabilizing soil quality, productivity, and 

water balance (Dayton, 1972; Ellison et al., 2005). For example, red 

maple and sugar maple, high-risk host tree species for ALB, are two of the 

most common tree species in many Northeastern forests in the United 

States (Lovett et al., 2006), and play a critical role in shaping the physical 

and chemical aspects of the soil (Mroz et al., 1985). Removal of 

foundation tree species can have dramatic effects on ecosystem function 

and stability. The physical characteristics of trees within forest 

ecosystems define forest structure and alter microclimates, while the leaf 

litter contributes substantially to ecosystem processes, such as nutrient 

cycling (Ellison et al., 2005). 

Host species that grow in riparian forest ecosystems offer a variety of 

ecosystem services specific to soil, including stability and nutrient cycling. 

The riparian zone is the interface between terrestrial and aquatic 

ecosystems through which water and materials move, and are often 

impacted by a change in soil conditions or vegetation (Knoepp and 

Clinton, 2009). The stability and nutrient cycling these ecosystems 

provide help to protect aquatic environments against excessive 

sedimentation, polluted surface runoff, and erosion. The overall health of 

riparian forest ecosystems is critical to maintaining good water quality 

and the health of stream ecosystems (Knoepp and Clinton, 2009). 

Soil erosion is the movement of soil particles by water, wind, or ice. 

Erosion is a natural process within ecosystems that removes and 

redistributes soil. However, anthropogenic activities (i.e., construction, 

agriculture) can accelerate erosion. A soil system is in equilibrium when 

soil erosion is in balance with the formation of new soil (Wall et al., 

2012). 

In forested sites on steep slopes, water is generally the most common 

driver of erosion. Erosion is usually infrequent in undisturbed forest soils 

because organic matter provides a protective layer on the soil surface, 

limiting the impact of raindrops, and allows water to infiltrate. The 

surface soil below the organic layer is generally porous, allowing water to 

infiltrate into and through the soil profile. Soil erosion can occur when the 

3. Soil Erosion 
40 III. Affected Environment 

surface soil is compacted, or when the loose surface soil and its protective 

layer of organic material are changed or removed, such as by disturbances 

associated with management activities. 

Forest soils may be compacted by grazing animals and by the roots of the 

trees themselves, but more noticeably by vehicles used for a range of 

mechanized forest operations. Soil compaction drastically reduces the 

number and size of pores that naturally occur throughout the soil. This 

reduces the exchange of gases and infiltration of water through the soil. 

Although compaction may allow surface soils to hold more water, it will 

tend to pool without soaking through to deeper layers of soil. In turn, 

runoff of surface water may increase, and tree growth may be reduced 

because of a reduced water supply, restricted root space, and poor 

aeration. In contrast, soil compaction may increase traction and, therefore, 

increase efficiency of vehicles moving on roads and tracks in the forest. 

D. Air 

This section provides a general overview of air quality as it pertains to the 

proposed action and alternatives. All site-specific assessments will 

include an analysis of local air quality, and may tier to this section of the 

programmatic EIS. 

Air quality is affected by two types of pollutants (primary and secondary), 

and greenhouse gases (GHG) that can pollute the air for human health, 

forest health, visibility, acid deposition, and climate change. Primary 

pollutants, such as particulate matter (PM), volatile organic compounds 

(VOCs), carbon monoxide (CO), nitrogen oxides (NOx), and sulfur oxides 

(SOx) that can affect human health, are directly generated from sources 

such as industrial facilities, cars and other mobile sources, and forest 

processes and activities, including fire (Stern, 1977). Some particulate 

matter such as dust, dirt, soot, or smoke are large enough to see while 

other particles are visible only with a microscope. Secondary pollutants, 

for example ozone (O3), are chemically transformed from primary 

pollutants such as VOCs and NOx (Stern, 1977). Forest fire emissions, 

when added to primary pollutants that affect human and forest health, 

contain carcinogenic air toxins such as acrolein, benzene, mercury, and 

formaldehyde (Langmann et al., 2009). Black carbon (BC) and GHGs, 

including carbon dioxide (CO2), methane (CH4), and nitrous oxide (N2O) 

affect climate. Pollutants such as sulfate (SO4), nitrate (NO3), and organic 

and elemental carbon (soot) reduce visibility (Stern, 1977). 

Atmospheric GHGs, such as CO2 and CH4, can trap solar energy and 

affect climatic conditions (EPA, 2012b). Elevated temperatures can lead 

to higher ozone levels (EPA, 2012b). GHG emissions in the United States 

increased 7.3 percent from 1990 to 2009, primarily due to emissions from 
III. Affected Environment 41

electricity generation and transportation transportation (EPA, 2012b) (table 

3–1). 

 

Table 3–1. GHG Emissions in the United States Allocated by Source*

Source 1990 2005 2008 2009 2010 2011 2012

Energy 5,260.1 6,243.5 6,071.1 5,674.6 5,860.6 5,712.9 5,498.9

Industrial 

processes 

316.1 334.9 335.9 287.8 324.6 342.9 334.4

Solvent and other 

product use 4.4 4.4 4.4 4.4 4.4 4.4 4.4

Agriculture 473.9 512.2 543.4 538.9 534.2 528.3 526.3

Land-use change 13.7 25.5 27.3 20.5 20.0 36.0 37.8

Waste 165.0 133.2 136.0 136.5 131.1 128.5 124.0

Total Emissions 6,233.2 7,253.8 7,118.1 6,662.9 6,874.7 6,753.0 6,525.6

Land use and 

forestry (sinks) (831.1) (1,030.7) (981.0) (961.6) (968.0) (980.3) (979.3)

Net Emissions 

(sources and 

sinks) 5,402.1 6,223.1 6,137.1 5,701.2 5,906.7 5,772.7 5,546.3

 

* Teragram (Tg), or million metric tons CO2 Eq. (Data source: (EPA, 2014b))

 

Different Federal, State, and local air regulatory agencies have created 

laws, rules, and regulations for control and reduction of air pollutants. 

Under the Clean Air Act (CAA), last amended in 1990, EPA set National 

Ambient Air Quality Standards (NAAQS) for pollutants considered 

harmful to public health and the environment (EPA, 2012a). NAAQS 

cover six criteria pollutants—sulfur dioxide (SO2), nitrogen dioxide 

(NO2), CO, ozone (O3), particulate matter (PM10 and PM2.5), and lead. 

EPA requires all States to develop attainment plans (State implementation 

plans (SIPs)) to improve air quality in nonattainment areas. 

Air quality monitoring data is collected and reviewed by EPA and State 

and local regulatory agencies, and is available to the public. This data is 

often published with respect to a local air quality index (AQI). AQI is a 

measurement of the level of pollutants in the atmosphere. An AQI above 

100 indicates that air quality conditions exceed human health standards, 

while values below 100 indicate pollutant levels are below air quality 

standards. An AQI that exceeds 100 suggests that air quality may be 

unhealthy for certain sensitive groups of people. 

In general, air quality is improving on a national scale, particularly 

through regulations and voluntary measures taken by industry to reduce 

emissions (EPA, 2012b). Cleaner cars, industries, and consumer products 
42 III. Affected Environment 

have contributed to cleaner air for much of the United States (EPA, 

2012b). Control programs for mobile sources and facilities (e.g., chemical 

plants, dry cleaners, coke ovens, and incinerators) are primarily 

responsible for these reductions (EPA, 2012b). Despite the downward 

trend in pollutant levels observed across the United States, numerous 

counties have reported nonattainment for one or more of the six criteria 

pollutants (EPA, 2013a). 

Trees have a beneficial effect on air quality by removing pollutants from 

the air, thus reducing human exposure to these substances and associated 

risk. Trees absorb CO2 from the atmosphere and release oxygen through 

photosynthesis (Shifley et al., 2012). Trees also absorb or intercept 

pollutant particles (PM10, e.g., dust, ash, dirt, pollen, and smoke) and 

gaseous pollutants (e.g., O3, NO2, and SO2) in the air, reducing human 

exposure to these substances and associated risks (Beckett et al., 1998; 

Nowak et al., 2000; Nowak et al., 2006; Tiwary et al., 2009). One 

estimate implies that a mature urban tree can intercept up to 50 pounds of 

particulates per year (Dwyer et al., 1992). In urban areas, ozone, SO2 and 

nitrogen oxide are some of the most common pollutants, all of which trees 

may absorb (Bell and Treshow, 2002). Trees protect human health by 

reducing pollutant exposure which, in turn, can diminish respiratory 

illnesses related to pollutant exposure (Kim and Bernstein, 2009). 

Forests can serve as a sink for GHG emissions, sequestering the gases that 

contribute to GHG levels (table 3–1). In 2009, estimates of the total 

carbon on forest land in the Northern States, including soil organic carbon, 

was 14,413 million dry tons, which represents about 32 percent stored on 

forest lands and soil in the United States (Shifley et al., 2012). In 2010, an 

estimated 12.4 billion tons of carbon was stored in Southern forests, 

within tree biomass, soil organic carbon, and understory plants above and 

below ground (Wear and Greis, 2013). Urban forests also contribute to 

carbon sequestration. In an early estimate, urban forests stored 

approximately 800 million tons of carbon, nearly 5 percent of live tree 

carbon storage in all U.S. forests (Dwyer et al., 1992). Fifty-year 

projections indicate the forest carbon pool in the South is 5 percent smaller 

than the pool in 2010 (a net emission of about 600 million tons) (Wear and 

Greis, 2013). Trees release stored carbon when burned or through the 

decay process. Conversion of trees to lumber does not immediately 

release carbon stores within the tree. 

Trees can have some negative effects on air quality as well. Trees release 

natural VOCs, which play a part in the formation of ozone and carbon 

monoxide (Beckett et al., 1998). These VOCs also can congregate with 

other particles in the atmosphere to create a haze over some stands of trees 

(Beckett et al., 1998). In a 2005 study, national VOC emissions from 

biogenic (natural) sources were larger than the VOC emissions from 

anthropogenic (human caused) sources, accounting for approximately 
III. Affected Environment 43

74 percent of VOC emissions (EPA, n.d.). Anthropogenic sources of 

VOCs are from industrial processes and manmade products, such as power 

plants, chemical production, solvents, vehicles, and other machinery 

(EPA, 2012b, n.d.). On a national level, anthropogenic VOC emissions 

have been declining (EPA, n.d.). 

In the Northeast region of the United States (comprised of 20 States), 

42 percent of the land cover is forested (Shifley et al., 2012). Projected 

urban development and other land use will shift the location of forested 

land cover, but overall coverage in the region is expected to remain stable 

in the near term (Shifley et al., 2012). The loss of trees in developed areas 

will affect local air quality; however, the greatest contribution to air 

quality improvements over the last decade is due to the reduction of 

mobile and industry emissions. “EPA expects air quality to continue to 

improve as recently adopted regulations are fully implemented and States 

work to meet current and recently revised national air quality standards.” 

(EPA, 2012b). Stricter air-quality regulations anticipated in coming years 

may add to the regulatory constraints on use of prescribed burning (Wear 

and Greis, 2013). 

E. Affected Biological Organisms 

All aquatic and terrestrial plant, vertebrate, and invertebrate species living 

in the environments that could support ALB are part of the affected 

environment. Changes in the composition of tree species can alter 

ecosystems, and can affect the species dependent upon them. Biological 

diversity, a term given to the variety of life and the natural patterns it 

forms, provides a large number of goods and services that sustain our 

lives, including food security, fresh air and water, energy, and biologicalbased products, such as wood products. In forests, high biological 

diversity enables these ecosystems to respond to external influences and 

recover from disturbances, while maintaining their ecosystem services, 

such as nutrient cycling, support to wildlife, and the purification of air and 

water (Shifley et al., 2012). Due to the importance of biological diversity 

in the health and sustainability of human populations, national and 

international programs and organizations work to preserve biological 

diversity, including the 150 Government-Member Convention on 

Biological Diversity, of which the United States is a member. Biological 

diversity is generally lower in suburban and urban forest communities and, 

in these areas, native species tend to be fragmented and small. Wildlife is 

usually limited to those species that are adapted to living near people, 

including raccoons, squirrels, deer, opossums, and a variety of birds. 

 

Domestic animals and pets also comprise a sector of animal life that 

cohabitates with people. Disturbances to wildlife through habitat 

destruction from development, traffic, and noise are common in developed 

communities. In contrast, less developed forest communities (e.g., State 
44 III. Affected Environment 

and national parks) sustain higher levels of biological diversity and harbor 

larger mammal species that tend to have a larger home range. In the 

Northeast region of the United States, forests support 780 known animal 

species (Shifley et al., 2012). Common mammals in the Northern forest 

include deer, black bear, porcupine, raccoon, and squirrel (Smith et al., 

2007). 

Migratory birds contribute to the biological diversity in the United States 

and bring enjoyment to millions of Americans. Neotropical migratory 

birds depend on forest stands for nesting and foraging (Donovan et al., 

1995; Suarez et al., 1997). Neotropical bird species vary in their habitat 

preference; some species depend on interior forest habitats, while others 

prefer edge habitats (Suarez et al., 1997; Thompson, 2005). In response to 

the importance of migratory birds, the United States established the 

Migratory Bird Treaty Act for the conservation of migratory birds and 

their habitats. The contiguous United States has four migratory game bird 

flyways: Atlantic Flyway, Central Flyway, Mississippi Flyway, and the 

Pacific Flyway. Managed by FWS and its partners, the goal for 

administering these flyways is to conserve migratory game bird species 

and allocate bird resources. Flyways are routes taken by a concentration 

of migratory birds between breeding and wintering areas. 

In the United States (not limited to the contiguous United States), there are 

over 1,400 (T&E) species federally listed through the Endangered Species 

Act. Approximately 60 federally listed species and their critical habitats 

co-occur with ALB-host trees and forest-dependent listed species in the 

Northeastern and mid-Atlantic areas (figure 4–2). Threatened species are 

plants and animals that are likely to become endangered throughout all or 

a significant portion of its range. 

Endangered species are those plants and animals that have become so rare 

that they are in danger of becoming extinct. The forest tree composition 

and its interaction with the air, water, and soil resources (discussed 

previously) affect the habitats in which T&E species live. T&E species 

are generally more sensitive to changes in their habitats. The threatened 

Virginia round-leaf birch, Betula uber (species of Betula are hosts to 

ALB), is found in only one location in Smythe County, Virginia. 

According to FWS, Virginia round-leaf birch is associated with second 

growth deciduous and mixed conifer/deciduous forests. Vegetation 

associated with the known extant population of the species includes oakpine and maple-beech-birch associations, with some tendencies to elmash-cottonwood associations perhaps because of the riparian setting 

(Garrison et al., 1977). In prior EAs, USDA–APHIS consulted with FWS 

on several T&E species that occur in areas where the Program is operating 

(see table 3–2 for a list of these T&E species). Several species are forestdependent, including the Indiana bat (Myotis sodalis) (FWS, 2007), and 
III. Affected Environment 45

the small whorled pogonia (Isotria medeoloides) (Mehrhoff, 1989). 

State species of concern, as well as federally listed plant species that are 

preferred hosts for ALB, occur in areas where ALB could become 

established (NHNHB, 2008). 

Table 3–2. Threatened or Endangered Species that Occur in Areas the Program is 

Operating 

Genus Species Common Name Type State Occurrence

Charadrius melodus Piping plover Bird MA

Cyprogenia stegaria Fanshell Aquatic mollusk OH

Epioblasma triquetra Snuffbox Aquatic mollusk OH

Isotria medeoloides Small whorled 

pogonia Terrestrial plant MA

Lampsilis abrupta Pink mucket pearly 

mussel Aquatic mollusk OH

Myotis sodalis Indiana bat Mammal OH

Myotis septentrionalis+ Northern long-eared 

bat Mammal MA, NY, OH

Plethobasus cyphyus Sheepnose Aquatic mollusk OH

Setophaga kirtlandii Kirtland's warbler Bird OH

Trifolium stoloniferum Running buffalo 

clover Terrestrial plant OH

Villosa fabalis Rayed bean Aquatic mollusk Ohio

* T&E species previously analyzed by APHIS for the ALB Program; APHIS consulted with FWS. 

+ Proposed for listing. 

 

Pollinators 

Pollination is a process of fertilization that occurs through the transference 

of pollen granules from the anthers (male parts) of a flower to the stigma 

(female parts) of the same or different flower. This ensures that a plant 

will produce a full-bodied fruit, and a full set of viable seeds. Pollen 

moves from flower to flower by wind, rain, and gravity, as well as by 

pollinating animals, such as birds, bees, bats, butterflies, moths, beetles, 

ants, and other animals. Pollinators use both pollen and nectar as food 

sources; some pollinators (e.g., honey bee) survive exclusively on pollen 

and nectar collected from flowers. 

In most terrestrial and aquatic environments, pollination and pollinators 

render vital ecological services. These services often have economic 

consequences; many agricultural crops rely on pollination to turn out the 

food on which humans and other animals depend for survival (AllenWardell et al., 1998). In fact, the ecological services that pollinators 

provide are necessary for the reproduction of over two-thirds of the 
46 III. Affected Environment 

world’s crop species, and 60 to 90 percent of the world’s flowering plants 

(Klein et al., 2007; Kremen et al., 2007). 

Reproduction of many flowering plants in the forest ecosystem is 

dependent upon insect pollinators (Coulson et al., 2005). Forest 

fragmentation caused by tree removal affects plant-animal interactions. 

Fragmentation introduces “edge” into a landscape—the changes in 

population or community structures that occur at the boundary of two 

habitats. The restricted size, discontinuity, and increased edge of 

fragments may impose many ecological and genetic effects on plants, both 

directly and indirectly, through pollination (Aizen and Feinsinger, 1994). 

Red maple, a regulated host of ALB in the United States, is an important 

early spring food resource for European honey bees (Apis mellifera) and 

other pollinators (Batra, 1985). When few other flowers are available, the 

red maple undergoes massive bloom between March and May, depending 

on elevation and latitude (Walters and Yawney, 1990). Boxelder and 

willow, both regulated hosts of ALB, are important food resources for 

mason bees (Osmia lignaria lignaria Say) during nest construction in the 

Northeastern United States and mid-Atlantic States (Kraemer and Favi, 

2005). 

Global pollinator decline has become an issue of concern for agricultural 

crop science. In North America and many parts of the world, the viability 

of multiple agricultural crops and broader ecosystems is threatened by 

unsustainable declines in the populations of honey bees, bumblebees, and 

other insect pollinators (NRC, 2007; Pettis and Delaplane, 2010). Colony 

collapse disorder (CCD) is a recent, pervasive syndrome affecting honey 

bee (Apis mellifera L.) colonies in the Northern Hemisphere, and is 

characterized by a sudden disappearance of adult honey bees from the 

hive. Multiple causes of CCD and general pollinator decline have been 

proposed, such as poor nutrition, pesticides, pathogens, parasites, and 

natural habitat degradation (Thompson, 2003; Desneux et al., 2007; Gill 

et al., 2012). 

F. Affected Economic, Social, and Cultural Factors 

Trees provide a range of products and support several industries within the 

United States. Tree species, such as those that are host trees for ALB, 

provide timber, maple syrup, and nursery trees, and generate income 

related to recreational activities. These industries operate predominantly 

in less developed forest communities, including uninhabited forest and 

forest recreational areas; these are important to the economies of many 

communities in the United States. The discussion below focuses on 

market and non-market values in the Northeastern United States as the 

number of ALB-host trees is higher in those areas when compared to other 

parts of the United States. In 2006, the Northern States wood products 

1. Market 

Factors 
III. Affected Environment 47

industry and the pulp and paper industry was estimated at $112 billion; 

primary wood products added $52 billion to the economy (Shifley et al., 

2012). Hardwood trees compose the majority of the industry in the North 

(Shifley et al., 2012). Economic returns on the market factors discussed 

below would be much greater when considering all States where ALBhost trees are present. 

The volume of sawtimber varies from State to State when making 

comparisons between Northeastern States where a significant amount of 

ALB host trees are present (table 3–3). Sawtimber refers to a growing 

stock tree containing at least a 12-foot saw log or 2 noncontiguous sawlogs 

8 feet or longer, free from defects. The percentage of hardwoods that are 

ALB host trees ranges from 24 percent for Rhode Island to 65 percent for 

Vermont. 

Timber from ALB-host trees has a variety of uses in roundwood products 

within these nine States. Roundwood products are logs, bolts, or chips cut 

from trees for industrial and nonindustrial uses (sawlogs, veneer logs, 

pulpwood, fuelwood, etc.) (FS, 2014). A majority of the volume from 

these tree species in these Northeastern States is used as sawlogs or as 

fuelwood (table 3–4) (APHIS, 2009). 

 Table 3–3. Volume of Sawtimber (ALB-Host Species) in Northeastern 

States and the State of Ohio—2012. 

State Net Volume*

(billion board feet)

% of Hardwood 

Volume at Risk

% of Total Volume

at Risk

Massachusetts 5.55 40% 22%

Connecticut 4.53 35% 30%

Maine 12.08 60% 21%

New Hampshire 7.49 49% 24%

Vermont 14.19 65% 38%

Rhode Island 0.43 24% 16%

New York 32.94 47% 35%

New Jersey 1.85 30% 15%

Pennsylvania 40.67 36% 34%

Ohio 17.33 36% 34%

* Net volume eguals gross volume less deductions for other defects that affect use for lumber.o. 

(Source: USDA–FS (2014), Northern Research Station (NRS-171:188)). 

 

This production of timber products translates into hundreds of millions of 

dollars in value in the Northeastern States. The total value of host species 

sawlogs for the eight States, as shown in table 3–5 (New Jersey is not 

included because stumpage prices were not available), was $171 million in 

2006. Production of veneer logs was worth an additional $720,000 in 

New York. Fuelwood production of host species was worth an additional 

2. Timber
48 III. Affected Environment 

$6.6 million in 2006. Although there is substantial fuelwood production 

in many States, the overall value of this product is less because the prices 

for fuelwood are significantly lower than many other products. Host 

species are also used for pulpwood, composite products, and other 

products in some Northeastern States, but are not valued here due to lack 

of price data (APHIS, 2009). 

Table 3–4. Volume of Roundwood Products (ALB-Host Species) in the Northeastern 

States—2006. 

Sawlogs

Veneer 

Logs Pulpwood

Composite 

Products

Fuelwood

Postpolespilings

Other 

Products

All 

Products

State MCF *

Connecticut 1,264 0 97 0 2,094 0 0 3,457

Maine 19,211 0 0 0 2,506 0 0 21,718

Massachusetts 950 0 0 0 7,095 0 0 8,044

New Hampshire 12,320 572 0 0 2,880 0 48 15,822

New Jersey 323 0 32 0 5,641 0 0 5,998

New York 28,902 588 21,879 1,440 41,805 0 2 94,612

Ohio** 60,272 478 23,538 695 No 

data 0 0 91,204

Pennsylvania 33,828 2,656 2,368 0 1307 298 554 41,014

Rhode Island 81 0 0 0 49 0 0 130

Vermont 12,360 0 0 0 7,598 0 0 19,956

TOTAL (not 

including Ohio) 109,239 3,816 24,376 1,440 70,975 298 604 210,751

 * = Thousand cubic feet 

** Data for Ohio is for 2010. (Source: USDA–FS, 2009)
III. Affected Environment 49

Table 3–5. Value of Select Timber Products for Species and State—2006. 

CT ME MA NH NY PA RI VT

(million dollars) 

Sawlogs 

(Total) 0.773 22.313 0.559 10.034 67.192 49.845 0.087 20.462

Ash 0.131 0.000 0.107 0.552 8.438 2.189 0.002 1.423

Sugar Maple 0.451 13.503 0.312 3.947 45.238 23.474 0.073 17.047

Red Maple 0.117 2.266 0.070 1.833 13.410 15.182 0.010 0.000

Yellow Birch 0.049 4.119 0.025 2.421 0.075 0 0.001 1.381

Other Birch 0.025 2.425 0.045 1.281 0.026 0 0.001 0.611

Elm 0 0 0 0 0.005 0 0 0

Veneer 

Logs 

(Total)

0 0 0 NA 0.7196 NA 0 0

Ash 0 0 0 NA 0.1323 NA 0 0

Sugar Maple 0 0 0 NA 0.3437 NA 0 0

Red Maple 0 0 0 NA 0.2391 NA 0 0

Yellow Birch 0 0 0 NA 0.0036 NA 0 0

Other Birch 0 0 0 NA 0.0007 NA 0 0

Elm 0 0 0 NA 0.0002 NA 0 0

TOTAL 0.773 22.313 0.559 10.034 67.912 40.845 0.087 20.462

Price data was not available for veneer logs and fuelwood in NH and PA. Source: (APHIS, 2009) 

 

The United States exports sawlogs to various countries. Figure 3–2 shows 

historical trends for the U.S. hardwood log exports for three ALB-host 

species. In 2014, log exports for birch, ash, maple were valued at $133 

million (USDA FAS, 2015). Historically, exports of maple logs have been 

the highest in value of the three species. These three ALB-host species 

comprise 21 percent of the value of U.S. hardwood log exports in 2014 

(USDA FAS, 2015). 

 
50 III. Affected Environment 

Figure 3–2. U.S. hardwood log exports for select ALB-host species, 1995–2014. 

 

In addition to timber production, ALB-host trees (e.g., maple) provide 

economic returns through maple syrup production. In 2013, total maple 

syrup production in the Northern region exceeded $131 million, with 3.1 

million gallons collected (table 3–6). Production in 2013 was greatest in 

Vermont with 1.4 million gallons ($49.4 million), followed by New York 

with 574,000 gallons ($25 million), and Maine with 560,000 gallons 

($17.9 million) (table 3–6). Other Northeastern States producing maple 

syrup include New Hampshire, Pennsylvania, Massachusetts, Connecticut, 

and Ohio, each of which produces 200,000 gallons or less annually. 

Demand in the United States has been increasing; currently only 0.4 

percent of tappable maples are being utilized, suggesting that economic 

returns and production could increase if demand continues (Farrell and 

Chabot, 2012). 

The tree nursery industry grows ALB-host trees. Between 2005 and 2009, 

producers in Pennsylvania sold 479,950 deciduous shade trees (not all 

were hosts of ALB) for $12.4 million; producers in New Jersey sold 

0

20,000

40,000

60,000

80,000

100,000

120,000

19951996199719981999200020012002200320042005200620072008200920102011201220132014

Value (Thousand dollars)

BIRCH

ASH

MAPLE

3. Maple 

Syrup 

4. Tree 

Nurseries
III. Affected Environment 51

Table 3–6. Maple Syrup Production 2010–13. 

Production (1,000 Gallons)

2010 2011 2012 2013

Vermont 890 1,140 750 1,480

New York 312 564 360 574

Maine 315 360 360 560

New Hampshire 87 120 76 124

Pennsylvania 54 128 96 134

Massachusetts 29 62 40 63

Connecticut 9 17 11 20

Ohio 65 125 100 155

TOTAL 1,761 2,516 1,793 3,110

Source: U.S. Crop Production, June 2014 USDA–NASS and USDA–NASS, New England Field 

Office, June 2014.}

 

300,119 trees for $22.2 million; producers in New York sold 363,008 trees 

for $12 million; and, producers in Ohio sold 218,341 trees for $16.6 

million (USDA, 2009). Within four Northeast States (Connecticut, New 

Jersey, New York, and Pennsylvania) there were 216 producers with over 

$56 million in gross sales in 2006 (USDA, 2009). 

Non-market factors relate to the benefits of ALB-host trees in residential 

and developed forest communities. Aesthetic values for residents and 

tourists; use values from recreation activities such as hiking, hunting, bird 

watching, and fishing; and ecosystem values from watershed services and 

carbon sequestration are examples of non-market factors related to urban 

and forested areas where ALB-host trees occur. Recreational activities, 

including fall foliage and wildlife viewing, hiking, and hunting generate 

revenue for many States. For example, visitors to Vermont spend $1.719 

billion in the State annually, of which approximately 27 percent ($460 

million) occurs during the fall season, from September through November 

(University of Vermont Tourism Data Center, 2012). Direct expenditures 

of visitors to Maine totaled $4.9 billion in 2012. Approximately 31 percent 

of trips to Maine occur during the fall season (October through 

November), which would account to around $1.5 billion in direct 

spending. Fifty-five percent of Maine visitors reported “beautiful scenery” 

as a motivation for their visit (DPA, 2013). 

Many factors affect tree value and benefits, including species composition, 

age distribution, condition, amount of canopy cover, and location. Urban 

trees provide valuable benefits to residents, including air temperature 

regulation, carbon sequestration, pollution reduction, stormwater runoff 

reduction, and lowering heating and cooling costs by serving as 

windbreaks, and casting shade resulting indirectly in the reduction of 

emissions from energy generation (Huang et al., 1987; Dwyer et al., 1992; 

5. Non-Market 

Factors 
52 III. Affected Environment 

McPherson and Simpson, 1995; Akbari et al., 1997; Simpson, 1998; 

McPherson, 2005). However, trees can also have the opposite effect and 

increase energy costs by casting shade on buildings in the winter, or 

blocking cooling winds in the summer (Nowak, 2002). 

In a literature review conducted by EPA, studies have found general 

increases of about 3 to 10 percent in residential property values associated 

with the presence of trees and vegetation on a property (EPA, 2008). 

Homeowners generally place a high value on their trees for shade, 

aesthetics, privacy, investment, and wildlife habitat, and are, consequently, 

concerned when this resource is threatened. 

The forestry industry supplies jobs to thousands of people, and is a 

dominant employer in some communities. In the Northern States, 441,000 

workers are employed in forest management, logging, forest products, and 

pulp and paper industries (Shifley et al., 2012). Other benefits of urban 

trees to society, that are more difficult to quantify, include increased job 

satisfaction, sentimental attachment, and improved child development 

(Kane and Kirwan, 2009). 

A person’s health is affected by the quality of the environment where they 

live. Access to greenness varies among demographic groups (Donovan 

et al., 2013). Trees are an important part of the natural environment, 

particularly in urban areas where they provide various health benefits to 

humans (Sarajevs, 2011), in addition to scenic views and environmental 

benefits. 

Trees also decrease human exposure to ultraviolet radiation through 

shading, which can reduce eye cataracts, and morbidity and mortality from 

skin cancer (Heisler and Grant, 2000; Heisler et al., 2003; Heisler, 2010; 

Grant et al., 2002). Greater tree-canopy cover has also been associated 

with improved birth outcomes, suggesting that trees may affect the health 

of a pregnant woman and reduce the risk of babies being born underweight 

(Donovan et al., 2011). Trees are part of urban green space that provides 

an environment conducive to physical activity. When there are green 

areas in the neighborhood, people tend to spend a greater amount of time 

outdoors and are more physically active (Humpel et al., 2002; Tzoulas 

et al., 2007). Children are less prone to become overweight when green 

space is available (Bell et al., 2008). Seniors in urban areas live longer 

when there are walkable green spaces along streets (Takano et al., 2002). 

Trees positively affect behavior and reduce crime (Kuo and Sullivan, 

2001a; Kuo and Sullivan, 2001b; Taylor et al., 2002) 

Trees affect the air, water, and soil in and around cultural and 

archaeological resources, as described in the respective sections above. 

Placement of trees around cultural, historic, and archaeological resources 

affect the aesthetic quality of the historic resource, and possibly the 

6. Social Factors

7. Cultural 

Factors 
III. Affected Environment 53

physical quality by providing wind buffer, shading, particulate adsorption, 

hydrological functions, and other protective attributes. Trees also can 

negatively affect cultural resources by damaging resources through falling 

limbs or trees, root growth, pollen production, and other natural 

phenomena. 

Trees themselves may also be part of the cultural traditions and heritage of 

various human groups. For example, wood fiber from ash trees is an 

important material used in baskets made by several Native American 

communities. Loss of ash trees in the United States from infestations of 

the emerald ash borer has affected the availability of this core resource to 

tribal communities (http://www.emeraldashborer.info/files/EABImpacts 

OnAmericanIndianCommunities.pdf). 

 
54 III. Affected Environment 

(This page is intentionally left blank.) 
IV. Environmental Consequences 55

IV. Environmental Consequences 

The purpose of this chapter is to discuss the potential impacts related to 

each of the proposed alternatives. Information presented in Chapter 3, 

Affected Environment, serves as the baseline for the evaluation of impacts 

to human health and the environment from the proposed alternatives. The 

potential impacts reflect those identified in the scoping notice and 

applicable comments received during the scoping process for this EIS, as 

well as the impacts evaluated in previous NEPA documentation for the 

ALB Eradication Program. Due to the programmatic nature of this EIS, 

the discussion in this chapter is a qualitative evaluation of the impacts of 

the proposed alternatives. A quantitative approach is used where possible 

and, in some cases, analysis for a specific geographic area is used as an 

example to relate the potential impacts of a proposed alternative. The 

information in this section, and the relevant appendices, are also 

applicable to a site-specific environmental analysis for an ALB outbreak. 

A. Program Alternatives 

a. Forest Resources 

Under the no action alternative, a worst-case scenario model, where ALB 

appears all at once in the Northeastern United States without mitigation in 

place, estimates a loss of 71 billion board feet of wood within 25 years, 

and death of all host trees within 60 years (Jacobson, et al., 2012. This 

would be a loss of about 45 percent of the forest (Jacobson et al., 2012), 

followed by succession with non-host plant and tree species that would 

result in a change in the current forest composition. This vegetation could 

replace ecological and economic voids left by the loss, for example, the 

growth of non-host trees may open other wood product markets. ALB has 

been in the United States for approximately 20 years. Under no 

eradication or mitigation measures, ALB populations would expand 

through natural spread. The spread of ALB through human-mediated 

pathways would also likely occur without quarantine restrictions. 

b. Environmental Resources 

(1) Water 

Tree loss and defoliation of ALB-infested trees have the potential to 

impact water quality. Many of the preferred hosts for ALB are tree 

species that occupy riparian and wetland areas. Riparian zones ensure 

high quality stream habitat for aquatic fauna. Loss of tree cover and 

density from ALB infestations can have negative impacts on streambank 

stabilization, water temperature, sediment loading, hydrology (increasing 

runoff), nutrient cycling, and contaminant removal in aquatic habitats 

1. Alternative 

1: No 

Action 
56 IV. Environmental Consequences 

(Jones et al., 2006; Lee et al., 2004; Wenger, 1999). The degradation in 

water quality from the loss of riparian function can also impact drinking 

water supplies, which has implications for human health. 

(2) Air 

Trees intercept pollutants from the air, reducing human exposure and 

associated risks, such as respiratory illnesses (Beckett et al., 1998; Nowak 

et al., 2000; Bell and Treshow, 2002; Nowak et al., 2006; Lovasi et al., 

2008; Kim and Bernstein, 2009; Tiwary et al., 2009; Donovan et al., 

2013). The estimated total annual air pollution removal for ozone, 

nitrogen oxides, sulfur oxides, carbon monoxide and particulate matter by 

urban trees in the United States was at 711,000 metric tons (Nowak et al., 

2006). One estimate implies that a mature urban tree can intercept up to 

50 pounds of particulates per year (Dwyer et al., 1992). Trees infested 

with ALB reach mortality within 10 to 15 years. During the declining 

stage, trees continue to intercept air pollutants and sequester carbon 

dioxide. However, stress to these trees decreases their ability to sequester 

carbon dioxide (Bréda et al., 2006). The loss of trees from ALB would 

cause a reduction in the interception of air pollutants and other air quality 

improvements; however, through natural succession or replanting with 

non-host trees, the air quality contributions would recover over time. 

Trees sequester the gases, including carbon dioxide, that contribute to 

GHG levels. During photosynthesis, plants and trees absorb carbon 

dioxide, store carbon above and below ground, and release oxygen as a 

byproduct. Trees release carbon back to the atmosphere through 

respiration, decomposition, and burning. Conversion of trees to lumber 

does not immediately release carbon stores within the tree. In 2009, 

estimates of the total carbon on forest land in the northern States, 

including soil organic carbon, was 14,413 million dry tons, which 

represents about 32 percent of the stored carbon on forest lands and soil in 

the United States (Shifley et al., 2012). In 2010, an estimated 12.4 billion 

tons of carbon was stored in southern forests within tree biomass, soil 

organic carbon, and understory plants above and below ground (Wear and 

Greis, 2013). Urban forests also contribute to carbon sequestration. In an 

early estimate, urban forests stored approximately 800 million tons of 

carbon, nearly 5 percent of live tree carbon storage in all U.S. forests 

(Dwyer et al., 1992). 

Climate change is the global shift in climate and weather from increased 

temperatures, mostly because of human activity. Carbon dioxide (CO2) 

and GHGs trap solar energy in the atmosphere, leading to increases in 

temperature. Trees, including ALB-host trees, store carbon and play a role 

in the reduction of CO2 in the atmosphere. The loss of trees to ALB 

reduces carbon sequestration, and the decomposition of dead trees release 

CO2 into the atmosphere. 
IV. Environmental Consequences 57

Climate change may affect the distribution of host trees and ALB. In a 

study of Northeastern U.S. forests, 36 out of 80 species assessed show the 

potential to shift their growing range approximately 62 miles (100 km) to 

the north, including 7 that could shift greater than 155 miles (250 km) 

(Iverson and Prasad, 1998). Sugar maple (Acer saccharum, 14.5%) and 

black cherry (Prunus serotina, 10.0%) would decline sharply, while oak 

and southern pines would expand northward. Temperature increases 

caused by climate change could affect ALB’s life cycle. In cold climates, 

ALB produces one generation every 2 years. In the areas of the 

United States where ALB has established, the insect typically produces 

one generation per year. The observance of one generation every 2 years 

may be possible if ALB establishes in colder climates in the United States. 

However, future increases in temperature may shift this to one generation 

per year. 

In the extreme scenario where ALB infests all host-tree species in the 

contiguous United States, (13.9 million metric tons (mt)) of CO2 would be 

released over time (depending on the rate of tree decomposition) to the 

atmosphere in the absence of any eradication (appendix D). If released in 

a single 1-year pulse, this would represent 37.8 percent of total global CO2 

annual emissions, based on the Year 2010 estimate of 9.1 Gt C/year (33.5 

gigatons (Gt) CO2) global emissions from industrial sources and an 

estimated total of 10.0 Gt C/year, including land-use change and 

deforestation (http:// co2now.org/Current-CO2/CO2-Now/global-carbonemissions.html). It is likely that this would have a significant but transient 

impact on overall climate change. 

Forests are capable of undergoing carbon recovery. Through sequestration 

in new growth, tree regeneration would recapture the equivalent of the 

CO2 released within 76 years (appendix D). Under improved management 

of tree health and tree cover, the level of CO2 sequestration in new tree 

growth could exceed initial losses due to tree mortality, and recapture fully 

all CO2 within as little as 43 years (appendix D). 

(3) Soil 

The loss of host trees in forests where ALB-host trees are dominant would 

alter soil nutrient cycling, especially in situations where a host tree serves 

a unique ecological function. For example, maple trees are considered a 

critical component in soil nitrification in the Northeastern United States; 

this loss could impact nitrogen retention and cycling in forested 

watersheds (Lovett and Rueth, 1999; Lovett and Mitchell, 2004). 

A decrease in maple stands from ALB infestation, would lead to higher 

nitrogen retention in soils and reduction in nitrogen transport into aquatic 

systems. The alteration of nitrogen cycling from the loss of maple would 

alter plant succession and diversity in terrestrial environments, as well as 
58 IV. Environmental Consequences 

impact aquatic ecosystems, that are dependent on higher nitrogen inputs. 

Impacts to nitrogen cycling in forested ecosystems have been reported 

with other defoliating invasive forest insects (Kizlinski et al., 2002; Lovett 

et al., 2002). Additional soil impacts may include increased soil erosion 

and increased soil temperature from the loss of trees, or foliage dieback 

resulting from repeated ALB infestations. 

c. Ecological Resources 

The dieback over successive seasons and loss of trees because of ALB 

infestation could result in a reduction of forested stands where ALB-host 

trees are dominant, and form canopy gaps in mixed forest habitats where 

ALB-host trees are present. The creation of canopy gaps and edge forest 

habitat can adversely impact those species that depend on contiguous 

blocks of forest for nesting and reproduction. In particular, neotropical 

migratory birds may be impacted by the loss or alteration of forest stands 

from ALB infestations. The species of birds impacted vary depending on 

whether it is dependent on interior forest habitats or edge habitats. Those 

species dependent on interior forest habitats may be negatively impacted 

due to the loss of habitat, while edge-dependent species may benefit. The 

loss of habitat for nesting and foraging impacts reproduction. For 

example, some neotropical migratory bird species could be negatively 

impacted by increased nest parasitization and predation from birds (e.g., 

brown-headed cowbird) in forested areas fragmented by the loss of ALBhost trees (Thurber et al., 1994; Donovan et al., 1995; Suarez et al., 1997; 

Thompson, 2005). 

Impacts to migratory birds vary by species and its habitat requirements, as 

well as the size of the ALB infestation (Chalfoun et al., 2002). Federal 

and State natural resource agencies, as well as non-governmental 

organizations, recognize the importance of habitats to migratory birds; 

these habitats are protected through bird conservation plans and bird 

conservation regions (BCRs). Two of those regions where ALB-preferred 

host plants are prevalent are the Atlantic Northern Forest and the 

Appalachian Mountain BCRs (FWS, 2006). The conservation plans for 

these areas identify threats to a range of bird species by invasive species, 

and forest habitat loss is a concern for multiple species. The added 

impacts from an ALB infestation on these regions increase the difficulty in 

implementing effective restoration and habitat protection plans. 

As with birds, other terrestrial fauna that depend on ALB-host trees and 

contiguous forested areas may be negatively impacted. Based on the 

spatial scale of forest loss, these impacts vary and are dependent upon the 

specific habitat requirements of the species. Canopy gaps or open areas 

resulting from an infestation may benefit some species while selecting 

against others that depend on ALB-host trees or depend on contiguous 

forest stands. Certain species may benefit from ALB-infested trees, such 
IV. Environmental Consequences 59

as secondary wood-boring pests and associated vertebrates that depend on 

snags for habitat and prey. 

A reduction in canopy cover or complete loss of ALB-host trees would 

favor understory vegetation. This would benefit species that depend on 

fragmented stands, and would create more open riparian areas for foraging 

and nesting (Bell and Whitmore, 2000). However, potential benefits to 

certain species from the loss of ALB-host trees are minor compared to 

widespread terrestrial and aquatic benefits related to the conservation of 

forested habitats. 

In addition to upland areas that could be impacted by ALB infestation, 

many of the preferred host trees for ALB commonly occur in riparian 

areas; impacts to these habitats can have negative impacts to bird species, 

as well as amphibians and other terrestrial, semi-aquatic and aquatic fauna 

(Naiman and Decamps, 1997; Petranka and Smith, 2005; Perkins and 

Hunter, 2006). Riparian areas are dynamic and complex habitats that serve 

as an interface between terrestrial and aquatic ecosystems, providing 

unique functions that, if significantly impacted, have negative 

consequences to a wide range of ecological processes within terrestrial and 

aquatic ecosystems (Naiman and Decamps, 1997). These areas provide 

biodiversity and habitat for numerous species, and regulate nutrient 

cycling and microclimate conditions in terrestrial and aquatic ecosystems. 

The loss of ALB-preferred host trees that are typical in these areas would 

impact the function of riparian zone habitats. In cases where ALB-host 

trees decline or are lost, other plant species grow that may contribute 

different or lower nutritive qualities, and alter decomposition rates. This 

can impact aquatic vertebrate and invertebrate species that exist in the 

particular habitat (Smock and MacGregor, 1988). The magnitude of 

impacts would vary, depending on the density and abundance of host trees 

and the plants and animals that rely on them for habitat. 

Figure 4–1 shows the types of direct and indirect impacts that could occur 

to representative nontarget aquatic and terrestrial species that are 

dependent on ALB-host trees. 
60 IV. Environmental Consequences 

Figure 4–1. Direct and indirect effects of ALB establishment to representative 

aquatic and terrestrial organisms.

d. Economic, Social, and Cultural Resources 

Both market and non-market impacts will occur as ALB-host trees are lost 

to the insect. Impacts will vary depending on the density of host trees in 

an area, the rate of spread of the insect, and the rate of decline of the trees. 

Hardwood trees compose the majority of the tree product industry in the 

north (Shifley, 2012), although not all of these are ALB hosts. 

Under a worst-case scenario model where ALB establishes in all 

Northeastern States, the loss of ALB-host trees in the Northeast region of 

the United States would result in a $20 billion loss of harvestable timber 

over 100 years, using a 3 percent discount rate to reflect the time value of 

money (Jacobson, 2012). This estimate takes into account the replacement 

of ALB-killed trees with other harvestable trees. 

The Northeastern region accounts for most of the maple syrup production 

in the United States and, based on the worst-case scenario where all 

maples would be lost, the maple syrup industry valued at $131 million in 
IV. Environmental Consequences 61

2013 would be lost (NASS, 2014). The tree nursery industry would lose 

part of their inventory; however, for those nurseries that carry non-host 

trees or are able to convert their inventory to non-host trees, an economic 

recovery may occur. Tree industries located within quarantine boundaries 

could incur additional costs due to restrictions (including treatments) 

imposed on the movement of ALB-host material. 

Countries listing ALB as a quarantine pest would restrict trade of 

commodities derived from host trees located within quarantine areas. 

Trees differ in the quality and utility of their wood. If replacement tree 

species do not meet export market needs, revenue from exports would 

decline. The United States may increase imports of timber and wood 

products derived from ALB-host trees should the country lose its domestic 

source. 

Other forest product industries could benefit from the loss of host trees. 

For example, host trees would supply material for firewood, biofuels, and 

pulpwood. However, a surplus in supply may cause a decrease in price for 

the commodity because the supply is greater than the demand. 

An immediate disappearance of ALB-host trees is unlikely to occur; rather, 

the natural spread of ALB is expected to occur slowly, as would the death 

of host trees, based on observations in the United States and elsewhere. 

As ALB-host trees disappear, other tree species may take their place, 

potentially sustaining the forestry industry, perhaps with the exception of 

the maple syrup industry. ALB is not the only problem affecting the 

health of Northeastern forests. Other invasive pests and diseases, as well 

as environmental stress (including that from climate change) and 

development pressures are affecting forest health throughout the United 

States. These stressors would be expected to affect forest recovery from 

ALB. 

Municipalities and landowners may experience a cost burden should trees 

on their property die from ALB infestation and require removal. The size 

of the tree and the number of trees requiring removal will influence cost. 

Indirect costs through tree removal include the impact to understory 

vegetation by the change in canopy closure and damage from removal 

activities. 

Research measuring the non-market value of trees is limited. The loss of 

ALB-host trees, particularly in areas dominated by these species, would 

change the composition and age of tree stands. Outdoor recreationists 

prefer areas with older trees (Scarpa et al., 2000; Englin et al., 2006), 

along with a preference for sites with more deciduous trees versus 

evergreen trees (Scarpa et al., 2000). Changes in forest composition, 

particularly with the loss of maples and other hardwood trees, would 

negatively affect the fall foliage tourism in the Northeastern United States. 
62 IV. Environmental Consequences 

Maples and other hardwood trees provide much of the brilliance in fall 

foliage. For example, in New Hampshire, fall foliage viewing contributes 

$292 million annually to the State’s economy (NH Dept. of Environ. 

Svcs.). Visitors to Vermont spend $1.719 billion in the State annually, of 

which, approximately 27.0 percent ($460 million) occurs during the fall 

season from September through November (University of Vermont 

Tourism Data Center, 2012). 

The loss of mature host trees would diminish the value and aesthetics of 

residential properties. Several studies show a correlation between both the 

presence and health of trees on property values (Holmes et al., 2006; Price 

et al., 2010). Holmes et al. (2006) studied the impact of hemlock wooly 

adelgid infestations on residential housing prices in northern New Jersey; 

a positive correlation was found between healthy hemlock trees and 

housing prices, and a negative correlation between infested hemlock trees 

and housing prices. In this study, unhealthy hemlock trees, up to 

0.62 miles (1 km) away from the property, had a negative effect on 

housing prices. 

A study in Portland, Oregon showed that street trees added an average 

$8,870 to sales prices, and reduced the time of housing on the market by 

1.7 days (Donovan and Butry, 2010). Other positive attributes trees 

impart to infrastructure include protection from stormwater runoff, 

temperature regulation, and electricity usage reduction, which would be 

affected by the loss of host trees. These effects would diminish over time 

as replacement trees mature. 

e. Human Health 

Under the no action alternative, ALB would continue to spread resulting in 

potential impacts to human health due to loss of trees. Tree loss in urban 

and rural residential areas was associated with declining cardiovascular 

health and lower-respiratory-tract illness in an ecological study evaluating 

the relationship between the presence of emerald ash borer and human 

health (Donovan et al., 2013). Tree loss from the spread of the emerald 

ash borer among 15 States was associated with 6.8 additional deaths per 

year per 100,000 adults over the 18-year period for the study (1990–2007) 

(Donovan et al., 2013). Tree loss contributes to risk factors for 

cardiovascular and respiratory diseases including stress, lack of physical 

activity, and poor air quality (Pope III et al., 2003; Everson-Rose and 

Lewis, 2005; Lucas and Platts-Mills, 2005; Donovan et al., 2013). 

Uncontrolled tree loss reduces the health benefits of trees discussed in 

chapter 3, and may cause negative effects on human health. 

Tree loss can cause stress and have psychological effects (Velarde et al., 

2007; Whitelaw et al., 2008). A comparison study on psychophysiological 

stress recovery and directed attention restoration of young adults, in 
IV. Environmental Consequences 63

natural and urban field settings, showed that natural settings with tree 

views have a positive impact on improved attention functioning, and 

lowered blood pressure levels (Hartig et al., 2003). This study found an 

increased positive affect and decreased anger for individuals in a natural 

reserve (Hartig et al., 2003). Another study analyzed survey results from 

953 individuals in 9 Swedish cities correlated increased participation in 

outdoor urban, open-green spaces with reduced stress levels (Grahn and 

Stigsdotter, 2003). (An open-green space is any open piece of 

undeveloped land with public access that is partly or completely covered 

with grass, trees, shrubs, or other vegetation.) Stress may also come from 

other sources, including decreased property values, and increased heating 

and cooling costs associated with tree loss (APHIS, 2013b). Tree loss can 

increase human exposure to ultraviolet (UV) radiation by removing the 

shading effects from trees that provide UV-radiation protection. UV 

radiation can negatively affect human health, primarily causing skin 

cancer and eye cataracts. UV radiation can also positively affect human 

health because it is essential for the synthesis of vitamin D. Vitamin D is 

required for bone health, and can reduce non-skin cancers. 

a. Forest Resources 

In one study, researchers modeled a worst-case scenario where ALB 

infests all host trees in the Northeastern United States all at once, with no 

eradication or mitigation programs in place. In this scenario, the Program 

would need to remove host trees from approximately 45 percent of the 

forested area (not including urban forests) (Jacobson et al., 2012). 

However, based on the historical introduction and spread of ALB, as well 

as APHIS’ response to the pest, the percentage of trees removed would be 

less, and would not occur throughout the United States or the Northeastern 

States all at once. APHIS has removed woodlot areas in Massachusetts 

and Ohio, resulting in the removal of hundreds of trees from one location. 

Urban forest cover would diminish if ALB-host trees dominate streets, 

parks, or residential plantings, as observed in cities and townships in 

Illinois, New York, New Jersey, Massachusetts, and Ohio. Regrowth of 

trees on woodlots through natural succession would take many years. 

Similarly, recovery of urban tree canopy cover will also take some time as 

replacement trees will likely be with immature non-host trees. 

Canopy cover influences understory plant composition, structure, and 

diversity (Oliver and Larson, 1996). Understory vegetation serves an 

important role in ecosystem health, biodiversity, and nutrient cycling, as 

well as contributions to soil and water quality in forests and wooded urban 

areas (Yarie, 1980; Oliver and Larson, 1996). Many vertebrate and 

invertebrate species depend on understory vegetation (Oliver and Larson, 

1996; Koide and Wu, 2003; Pineda et al., 2005). The removal of trees 

infested with ALB would reduce the canopy cover, creating an 

environment favorable to shade-intolerant vegetation. 

2. Alternative 

2: Removal of 

Infested 

Trees
64 IV. Environmental Consequences 

The process of removing trees can damage surrounding vegetation. Felled 

trees, vehicles, and other tree removal equipment can compress vegetation 

and soil. The introduction of weeds and invasive plant species on 

equipment could alter the vegetative understory. In some instances, the 

Program applies an herbicide to stumps to prevent sprouting, although the 

preference is to grind the stumps in place. During recent ALB eradication 

efforts, greater than 75 percent of the stumps were removed rather than 

being treated with an herbicide. While herbicide application is directly on 

the stump surface, and according to label instructions, damage to nearby 

vegetation could occur from drift or runoff. 

b. Environmental Resources 

(1) Water 

The removal of infested trees near aquatic resources can impact water 

quality. In particular, the movement of soil into aquatic resources (rivers, 

lakes, and other bodies of water) can result in sedimentation, excessive 

nutrients (eutrophication), increased turbidity or cloudiness, and alteration 

of stream flow. In addition, tree removal adjacent to aquatic resources can 

reduce shading, which is important in maintaining water temperature. 

Degradation of water quality due to sedimentation can result in negative 

effects to aquatic organisms through direct or indirect impacts to fish, 

aquatic insects, and crustaceans, such as freshwater mussels and crayfish 

(Richter et al., 1997; Henley et al., 2000). The risk to soil quality and 

aquatic resources from erosion, due to tree removal, can be reduced by the 

implementation of timber BMPs (Aust and Blinn, 2004). 

(2) Air 

As described in the environmental consequences for the no action 

alternative, the loss of trees to ALB infestation could affect air quality and 

contribute to climate change. The degree to which air quality and climate 

change are affected depends on the human-facilitated contribution of air 

pollutants, as well as the number of trees infested with ALB. 

Trees infested with ALB reach mortality within 10 to 15 years. During the 

declining stage, trees continue to intercept air pollutants and sequester 

CO2; however, stress to these trees decreases their ability to sequester 

carbon dioxide (Bréda et al., 2006). Removal of trees prior to mortality 

reduces the interception of air pollutants and sequestration of carbon 

dioxide. Trees release stored carbon during decomposition or burning. 

Under this alternative, the Program removes and chips infested trees. 

Wood chips decompose at a faster rate than intact woody material, 

resulting in a faster release of carbon dioxide (McPherson and Simpson, 

1999). Replacement of trees through succession or planting would restore 
IV. Environmental Consequences 65

air quality attributes to the area; however, this would occur slowly over 

time and would vary depending on the types of species that may regrow in 

areas. 

The loss of trees through the removal of ALB-infested trees by the 

Program would temporarily affect the local carbon sequestration. For 

example, the potential total CO2 release estimate from trees and soil at five 

ALB eradication sites active in 2012 is 20,187 mt (appendix D). These 

levels are below the CEQ reference level of 25,000 mt for all GHGs; other 

GHGs (methane, nitrous oxide, hydro-fluorocarbons, perfluorocarbons, 

and sulfur hexafluoride) were not included in this study. The restoration 

of vegetation and trees through natural succession or intentional planting 

reduce the contribution of CO2 to the atmosphere from the removal of 

trees by the Program. 

In urban areas, where trees now shade buildings, tree removal under ALB 

eradication would increase energy requirements and emissions of GHGs 

from power plants to compensate for increased heating in winter and air 

conditioning in summer. For example, in New York City, trees are 

estimated to reduce energy costs from residential buildings by $11.2 

million annually based on 2002 energy costs. Trees also provide an 

additional $167,000 in value per year by reducing the amount of carbon 

released by the city’s fossil-fuel based power plants (a net reduction of 

9,100 tons of carbon emissions) (Nowak et al., 2007). 

Several U.S. cities have greening programs (e.g., Boston 

(http://www.cityofboston.gov/parks/streettrees/growboston.asp) and 

Baltimore (http://www.baltimoretreetrust.org/) aimed at increasing tree 

cover to achieve benefits trees provide, including improved air quality, 

sequestration of carbon, reduced energy consumption, and flooding 

control (Nowak et al., 2007; Nowak et al., 2010). The removal of infested 

trees prior to mortality in urban areas and adjacent forest would negatively 

affect these benefits. However, infested trees weaken over time and 

eventually fail to benefit the urban environment; rather, they may become 

a fall hazard. In addition, leaving infested trees leads to additional tree 

infestations due to ALB spreading resulting in more tree loss. 

The combustion of gasoline and diesel fuel in machinery used to remove 

and chip trees release air pollutants and GHGs (McPherson and Simpson, 

1999). Tree removal may also contribute wood and soil particulates to the 

air during wood chipping and disturbance of soil by machinery during 

removal. Estimates of release for these activities are scarce, including 

data from ALB Eradication Program activities. Emissions from tree 

removal and chipping activities would have the greatest impact to air 

quality in urban areas where air quality may already be impacted. 
66 IV. Environmental Consequences 

(3) Soil 

Soil quality impacts under this alternative (Alternative 2: Removal of 

Infested Trees) would be similar or, in some cases, more significant than 

those described under the no action alternative. Under this alternative 

infested trees would be removed, which could result in physical and 

chemical impacts to soils, especially in areas where soils are vulnerable to 

erosion. The removal of only infested trees without addressing host trees 

nearby at risk of infestation could also allow ALB to continue to spread, 

resulting in additional tree removals in areas where soils may be 

susceptible to erosion. 

c. Ecological Resources 

The removal of infested trees would have impacts to ecological resources 

similar to those described under the no action alternative, although at a 

potentially slower rate in the long term. Initially the rate of tree loss under 

this alternative would be greater in the infested areas compared to the no 

action alternative since trees would be removed more quickly than through 

natural loss from ALB. In cases where infested trees are removed, some 

of the impacts previously noted could occur. However, the expansion of 

ALB resulting from not removing high-risk host trees within the known 

dispersal range of ALB would leave those trees vulnerable, and 

infestations would continue to occur, resulting in additional removal of 

infested trees. 

Herbicide Use 

APHIS uses herbicides when there are limitations to physical removal of 

stumps. The limitations include those areas that are inaccessible to 

equipment used for stump grinding, and those areas that are sensitive to 

erosion or compaction. 

The Program uses the herbicide triclopyr by spraying or painting the root 

collar area (the sides of the stump) and the outer portion of the cut surface, 

including the cambium (thin layer of generative tissue lying between the 

bark and the wood of a stem), until thoroughly wet, but not to runoff. 

Foliar applications of triclopyr mixed with two other herbicides, imazapyr 

and metsulfuron-methyl, would be applied to sprouting foliage from 

stumps that remain after tree removal to prevent regrowth. 

Triclopyr triethylamine salt (TEA) toxicity to terrestrial wildlife is 

considered low. Toxicity to avian species is low for triclopyr TEA with 

oral and dietary median lethal toxicity values greater than the highest test 

concentrations tested (EPA, 1998; Durkin, 2003). Chronic toxicity to 

birds is also expected to be low with reproductive toxicity. The no 

observable effect levels (NOEL) are 100 and 500 parts per million (ppm) 
IV. Environmental Consequences 67

for the mallard and bobwhite quail, respectively, when exposed to 

triclopyr acid (EPA, 1998). Available avian toxicity data for triclopyr 

butoxyethyl ester (BEE), another triclopyr product available for use by the 

Program, demonstrates slight toxicity, with median lethal dose values 

ranging from 735 to 849 mg/kg for the bobwhite quail (EPA, 1998). 

These values are well above any residues that would occur due to Program 

applications. Triclopyr TEA is not toxic to honey bees based on acute 

contact studies (EPA, 1998). Triclopyr TEA does exhibit toxicity to some 

terrestrial plants based on results from seedling emergence, germination, 

and vegetative vigor studies. The primary degradation product of triclopyr 

TEA, triclopyr acid, is similar in toxicity to terrestrial nontarget organisms 

based on the available toxicity data. 

TEA toxicity to aquatic organisms is low for fish and aquatic 

invertebrates. Available acute fish toxicity data demonstrates median 

lethal concentrations greater than 100 mg/L for Garlon® 3A and technical 

triclopyr TEA (Wan et al., 1987; EPA, 2014a). Triclopyr TEA is not 

considered toxic to aquatic invertebrates in freshwater and marine 

environments, with toxicity values exceeding 300 mg/L. Chronic toxicity 

to fish and aquatic invertebrates is also low with chronic toxicity no 

observable effects concentration (NOEC) ranging from approximately 

80 mg/L to greater than 100 mg/L, depending on the test organism and 

endpoint. Although Triclopyr BEE may be toxic to aquatic invertebrates 

and fish, they will not be exposed to levels that could result in adverse 

effects from applications made by the Program. The primary metabolite 

of triclopyr TEA and BEE, triclopyr acid, is not considered toxic to 

aquatic organisms, based on available toxicity data (EPA, 1998, 2014a). 

For foliar treatments, Garlon® 3A is proposed for use as a mixture with 

the active ingredients imazapyr and metsulfuron-methyl. Imazapyr is an 

imidazolinone herbicide while metsulfuron-methyl is a sulfonylurea 

herbicide, with both products used as a mixture with triclopyr in the 

control of woody vegetation. 

The toxicity of imazapyr and metsulfuron-methyl is considered low for 

mammals. The formulation containing metsulfuron-methyl, Escort® XP, 

is not considered toxic to mammals via inhalation, dermal, and oral 

exposures. All toxicity values were reported as greater than the highest 

test concentration. In addition, metsulfuron-methyl is not considered to be 

carcinogenic, nor has it been shown to be a reproductive, teratogenic, or 

developmental hazard (Klotzback and Durkin, 2004). Escort® XP is 

considered a slight eye irritant, but is not considered a skin irritant or 

sensitizer. The other herbicide in the mixture, Arsenal®, containing the 

active ingredient imazapyr, has a similar mammalian toxicity profile to 

metsulfuron-methyl, and is considered practically nontoxic in acute 

inhalation, dermal, and oral exposures. Imazapyr is not considered a 
68 IV. Environmental Consequences 

carcinogen or mutagen, and is not known to be a reproductive, teratogenic, 

or developmental hazard (Durkin and Follansbee, 2004). 

The toxicity of imazapyr and metsulfuron-methyl is low to all nontarget 

organisms, with the exception of some aquatic and terrestrial plants. 

Neither product is considered toxic to mammals, birds, or terrestrial 

invertebrates (Durkin and Follansbee, 2004; Klotzback and Durkin, 2004; 

EPA, 2014a). Toxicity to fish and aquatic invertebrates is very low, with 

median lethal acute concentrations typically exceeding 100 mg/L for both 

chemicals (Durkin and Follansbee, 2004; Klotzback and Durkin, 2004; 

EPA, 2014a). Chronic toxicity to fish and aquatic invertebrates is also 

considered low, based on the available NOECs that were reported from 

standardized toxicity studies. 

Exposure to terrestrial and aquatic nontarget organisms is expected to be 

minimal from each proposed formulation and mix. Significant drift or 

runoff is not expected as applications are not broadcast applied, but are 

made using either a backpack sprayer to deliver a coarse droplet size, or 

by brushing the material on individual stumps and associated sprouting 

vegetation. The low probability of offsite transport for any of the products 

results in very low exposure to most nontarget organisms. The low 

probability of exposure and the favorable available effects data 

demonstrate that all products have a very low risk of causing adverse 

ecological risk (see appendix E). Risk to nontarget organisms is greatest 

for plants as they are the most sensitive group to each application; 

however, the application methods and label directions minimize impacts 

to terrestrial plants, restricting potential harm to those plants that are 

immediately adjacent to treated stumps or sprouts. Exposure in aquatic 

systems is not expected to occur at levels that could result in any direct 

impacts to aquatic plants, or at levels that would suggest indirect impacts 

to aquatic organisms that depend on aquatic plants as a food source or as 

habitat. (Appendix E provides the risk assessment for herbicides the 

Program proposes to use.) 

d. Economic, Social, and Cultural Resources 

As described in the environmental consequences section for the no action 

alternative, ALB-host trees are important to the forestry products industry 

and their loss, particularly on forested lands, would result in negative 

economic impacts. The Program removes and chips infested trees, making 

them unavailable to the timber and forestry products industry. ALB larvae 

create tunnels or galleries inside the tree, damaging the structural integrity 

of the wood. It is possible that lightly infested trees could have 

salvageable wood for timber and other end-use products; however, the 

Program does not allow the diversion of infested trees from chipping to 

saw mills because of the risk of spreading ALB. 
IV. Environmental Consequences 69

One maple tree can produce 10 to 60 gallons of sap for maple syrup in one 

season, depending on the tree (including size), weather conditions, length 

of sap season, and the method of collecting sap 

(http://maple.dnr.cornell.edu/index.html). Maple syrup producers can 

absorb a loss of a percentage of maple trees but, depending on the size of 

the producer, there is a threshold where economic loss would shut down 

the business. APHIS does not recommend replacing maple trees with 

ALB-host trees in an area infested with ALB; therefore, replanting would 

not be the best option for maple producers. 

The impact to municipalities and landowners is similar to those described 

under the environmental consequences for the no action alternative. 

However, the Program will remove infested trees rather than leaving them 

to die in place. In the early stages of infestation, trees can appear healthy 

and continue to provide the aesthetic qualities and other benefits. Trees 

decline and die at different rates, depending on type of tree, its size, the 

population of ALB, exposure to other stressors, and other factors. 

Symptoms occur in approximately 3 to 4 years after infestation, and tree 

death can occur in 10 to 15 years, depending on site conditions. The 

Program would remove trees at all stages of infestation. This alternative 

would slow the spread of ALB; however, due to the difficulties in 

identifying infested trees, it is likely that some infested trees would be 

missed and, therefore, ALB would spread. 

In general, damage to road surfaces may occur in areas with high traffic 

volume, frequent use by heavy vehicles, and weather events such as freezethaw cycles. Tree roots can also distress sidewalks and roads causing 

cracks, vaults, and faults. The quality of the road construction materials 

and the road’s design will affect the road’s rate of decline (Charlier et al., 

2009). The removal of trees in the ALB eradication program may 

contribute to road deterioration in some areas through the use of heavy 

equipment to remove the trees; however, APHIS expects the contribution 

to be minimal given the existing traffic volumes on roadways and the short 

duration heavy equipment used in the Program will travel on roads in the 

impacted area. Roads consist of several layers of different porosity which 

aid in water permeability (Charlier et al., 2009). Road design for porosity 

is usually independent of the water uptake from nearby trees. Water 

accumulation can cause surface deterioration (Randup et al, 2001 cites 

Army Corps of Engineering; Charlier et al., 2009). The removal of trees in 

the ALB Program may cause an initial increase in water availability in the 

area through the reduction of plant biomass available to uptake water 

(Randrup et al., 2001). After tree removal, the ALB Eradication Program 

plants a groundcover to reduce erosion. This site stabilization as well as 

any replacement of trees will remediate water runoff. 
70 IV. Environmental Consequences 

e. Human Health 

Under this alternative, the overall rate of tree loss from ALB infestation is 

expected to be reduced from the no action alternative because the Program 

removes infested trees, which reduces the ALB population and spread to 

other host trees. However, tree removal activities would initially cause an 

increase in tree loss, compared to the no action alternative where infested 

trees remain in place to die from ALB-infestation. In the short term, tree 

removal activities may result in increased noise levels (from use of 

mechanical equipment and increased traffic), increased stress (from 

decreased property values), increased cooling and heating costs, and other 

localized negative human health consequences from a lack of trees, (as 

discussed in the no action alternative) to the general public living in the 

infested areas. Emissions and particulates from tree removal activities 

may cause a short-term disturbance in air quality, particularly for 

sensitive groups. In the long term, the negative human health 

consequences could be less than those compared to the no action 

alternative because the overall tree loss is expected to decrease to some 

degree. 

(1) Herbicide Use 

APHIS evaluated the risk to workers and the general public from the 

Program use of the three herbicides (appendix E). Two triclopyr 

formulations, Garlon® 3A (active ingredient is TEA) and Pathfinder® II 

(active ingredient is BEE), for the treatment of stumps were analyzed. 

Pathfinder® II is used to treat the bark instead of direct application to cut 

areas of the stem. Minor foliar applications of Garlon® 3A mixed with 

two other herbicides, imazapyr and metsulfuron-methyl, are used to treat 

sprouting foliage from stumps that have been removed as part of the 

eradication program. The applications are made by hand either by 

brushing undiluted material on the stump or directly spraying stumps 

and/or sprouting foliage using a backpack sprayer. The TEA formulation 

can cause significant eye irritation, but has low acute inhalation and 

dermal toxicity. Acute oral median lethal concentrations range from 

approximately 600 to 1,000 mg/kg, suggesting low to moderate toxicity 

(Durkin, 2003).

Long-term toxicity studies have shown that triclopyr TEA is not a 

carcinogen or mutagen, and that toxicity in developmental and 

reproductive studies primarily occurs at high doses, and at levels that are 

also maternally toxic (EPA, 1998). The concentrations at which these 

effects have been reported would not occur under normal program uses. 

The other proposed BEE formulation, Pathfinder® II, can cause slight 

temporary eye irritation during application, and some skin irritation under 

prolonged exposure. Acute oral median lethal concentrations are 
IV. Environmental Consequences 71

1,000 mg/kg, with acute inhalation and dermal toxicity median lethality 

values greater than the highest test concentration, suggesting low acute 

mammalian toxicity under various exposure pathways. Triclopyr BEE is 

not considered carcinogenic or mutagenic and, in cases where 

developmental and reproductive studies demonstrate effects, doses were at 

levels considered maternally toxic. The concentrations at which these 

effects have been reported would not occur under normal program uses.

TEA breaks down in soil (~12 days) to triclopyr acid, and to a lesser 

extent, triethanolamine. Triclopyr BEE has low water solubility, and 

adsorbs more strongly to soil when compared to the amine. Triclopyr 

BEE also breaks down quickly to triclopyr acid in soil and water, with 

hydrolysis half-lives of less than 1 day. Imazapyr degradation and 

dissipation half-lives are variable, ranging from approximately 25 days to 

greater than 300 days. Metsulfuron-methyl half-lives in soil range from 

17 to 180 days. The proposed use pattern and available chemical data for 

each herbicide suggests that these products will not bioconcentrate or 

bioaccumulate in the environment (appendix E). The use of these 

herbicides may negatively affect sensitive individuals or those who 

inadvertently become exposed; however, these herbicides are relatively 

short lived in the environment so any health effects from herbicide use are 

likely to be temporary. 

The human health risk assessment results show that using triclopyr, or 

using triclopyr and mixing it with imazapyr and metsulfuron-methyl to 

control ALB populations should pose minimal risks to human health for 

workers and the general public (appendix E). 

In addition, notification to landowners also occurs in the case of any 

chemical treatments that may be used to kill stumps (herbicides), as well 

as other label risk reduction requirements for herbicide use that are 

designed to protect workers, the general public, and the environment. 

a. Forest Resources 

Host trees within a 1⁄2-mile radius of an infested tree are at risk of ALB 

infestation because they are within the dispersal range of the beetle. 

Under the full host removal alternative, removal by the Program of both 

infested trees and high-risk host trees would occur. Depending on the 

density of host trees in the quarantine area, this will likely translate to the 

removal of a larger population of trees compared to the no action 

alternative and two of the eradication alternatives. The intent is to protect 

urban and rural forests from ALB through the removal of host trees 

potentially infested with ALB but not at detectable levels. 

The impacts to forestry resources, as described in alternative 2, could also 

occur under this alternative. Trees infested with ALB typically die within 

3. Alternative 

3: Full Host 

Removal 
72 IV. Environmental Consequences 

10 to 15 years. Full host removal by the Program prematurely eliminates 

the ecological services the trees provide. Host trees provide habitats and 

food for wildlife, and contribute to nutrient and water cycles. The longterm benefits of full host removal in a quarantine area include the 

reduction in the beetle’s spread rate and eradication of the beetle. 

Removal of host trees classified as invasive in the United States would be 

beneficial. For example, several Northeastern States designate the 

Norway maple as an invasive species. 

b. Environmental Resources 

(1) Water 

Impacts to water quality from the selection of alternative 3 would be 

increased when compared to alternative 2. The removal of infested trees 

and high-risk host trees would result in soil sediment runoff to aquatic 

areas. Current water quality data for rivers and streams in the United 

States lists sediments as the second leading cause of impairment under 

Section 303(d) of the CWA (EPA, 2014c). Sediments are the ninth 

leading cause of impairment in lakes, reservoirs, and ponds. The source of 

sediments causing impairment in waterways varies; however, silviculturerelated activities, such as harvesting and forest road construction, are 

contributing factors. These impacts would be more likely in forested areas 

where large numbers of trees may be removed as a result of an ALB 

infestation. Replanting groundcover and following silviculture BMPs will 

protect soils vulnerable to erosion, reducing potential for impacts to water 

quality. 

(2) Air 

Under this alternative, the potential removal of a greater number of trees 

could cause greater impact to air resources compared to those described in 

Alternative 2—Removal of Infested Trees. Loss of all host-tree species 

across the contiguous United States is estimated to release 13.9 million mt 

of CO2 to the atmosphere (appendix D). This large pulse of CO2 to the 

atmosphere is unlikely as ALB spreads slowly, trees become infested at 

different times, and trees die at different rates. Large urban and forested 

areas under quarantine with a high density of ALB-host trees would see 

impacts on the local air quality. 

Forest areas are typically a mixture of tree species. For example, in the 

Northeastern United States, forests are typically a mixture of hardwood 

and hardwood-conifer (Shifley et al., 2012); full host removal would not 

result in complete deforestation of an area. Emissions from tree removal 

and chipping activities could result in some localized impacts to air 
IV. Environmental Consequences 73

quality, but these would be reduced over time as eradication efforts are 

implemented. 

(3) Soil 

This alternative would result in potentially greater impacts to soil quality, 

compared to the other eradication alternatives, because both infested trees 

and high-risk host trees would be removed. Impacts to soil would be 

greatest where large numbers of trees are removed from a concentrated 

area. 

Changes in soil temperature and moisture, as well as soil erosion and loss 

of nutrients in areas, can impact the ability of a forest to regenerate 

(Ballard, 2000). These impacts are more prevalent in cases where clearcutting is being used for forest harvesting; however, that type of removal 

is not likely to occur in the Program because other non-ALB-host trees 

would be left standing. 

Compaction from the use of heavy equipment may result in increased soil 

bulk density values that may limit regrowth of vegetation in areas where 

trees are removed. These physical impacts to soil may result in increased 

erosion of soil from wind and rain, both during and after tree removal. 

Seeding areas with groundcover reduces these types of soil quality 

impacts. 

c. Ecological Resources 

Overall impacts to ecological resources, under alternative 3, are expected 

to be less than those described under the no action alternative and 

alternative 2 because of the prevention of ALB dispersal to non-infested 

areas within the United States. Impacts on a local level may be more 

significant to ecological resources because removal of infested and highrisk host trees may result in some fragmentation, as well as a reduction in 

tree density within riparian zones. While these impacts would be more 

localized, compared to the loss of trees resulting from the no action 

alternative, the impacts would be more immediate as trees would be 

removed at a more rapid rate than if they were lost to ALB. The extent of 

ecological impacts would be dependent upon the size of the infestation 

and the predominance of ALB-host trees in natural and urban areas. The 

density of host trees will likely be greater in forest areas than urban 

settings, potentially causing greater impacts to ecological resources in 

forest areas. 

d. Economic, Social, and Cultural Resources 

Full host tree removal would result in the removal of more trees within the 

quarantine area compared to the other three eradication alternatives. The 
74 IV. Environmental Consequences 

extent of the impact depends on the density of host trees within the area 

and the intended use of the host trees. Significant impacts could occur if a 

quarantine area overlaps with a woodlot, forest, tree plantation, maple 

syrup production area, or other commercial forestry area dependent upon 

hardwood trees that are hosts to ALB. 

Urban areas predominantly planted with ALB-host trees would see a 

reduction in tree cover. This may result in an initial reduction in aesthetic 

qualities of the landscape, as well as the ecosystem services trees provide; 

this includes reduction in water and sediment runoff, interception of air 

pollutants, and buffering from solar radiation and wind. 

The program restores tree removal areas through grading and planting of 

groundcover and in the case of woodlots, allowing restoration through 

natural succession leading toward recovery of resources. Municipalities 

and landowners could incur a cost depending on restoration activities such 

as the planting of non-host trees or other vegetation. Suppression in the 

tourism industry may occur if a high proportion of the ecological resource 

(e.g., park, community, etc.) falls within the quarantine area. Full host 

removal is expected to have less of an economic impact compared to the 

no action alternative because this alternative slows the spread of ALB and 

leads to eradication. Compared to Alternative 2—Removal of Infested 

Trees, Alternative 3—Full Host Removal is expected to produce greater 

short-term economic impacts because of the potential removal of more 

host trees. However, in the long- term, it is expected that impacts would 

be reduced because of the protection of forest resources.

e. Human Health 

Alternative 3 would result in increased efficacy to eradicate ALB, and 

prevent its spread to new areas. In the short term, tree loss would occur 

from removal of infested and surrounding host trees. These removals 

could result in localized negative human impacts related to tree loss, such 

as increased stress, reduced air quality, and so on, as discussed in the no 

action alternative. However, in the long term, the rate and spread of tree 

loss due to ALB would be reduced, minimizing human health related 

impacts. 

The potential for human exposure and risk to herbicide use is the greatest 

under this alternative because more trees are being removed and herbicide 

use would be expected to increase. Human health risks are still expected 

to be low based on the herbicide risk assessment included in appendix E of 

this EIS. In some cases, stumps may be left to allow for regrowth in areas 

where ALB reinfestation is not likely to occur. When stump treatment is 

needed, grinding is the preferred method over herbicide treatments; 

therefore, while there would be an increase in herbicide use, it would not 

necessarily be proportional to the increase in the number of trees removed. 
IV. Environmental Consequences 75

a. Forest Resources

Under this alternative, the Program removes host trees infested with ALB, 

and treats the high-risk host trees located within a 1⁄2-mile radius of 

infested trees with the insecticide imidacloprid. The Program applies 

insecticide treatments in the fall, spring, and early summer prior to and 

during the adult emergence period. The impact to forestry resources from 

this alternative include the impacts described under forestry resources for 

Alternative 2: Removal of Infested Trees. 

In addition to those impacts, the treatment of high-risk host trees with 

imidacloprid within a 1⁄2-mile radius of infested trees may affect beneficial 

insects (e.g., honey bees) associated with host trees. Research is ongoing 

to understand these interactions and their impacts on tree health and 

reproduction. Research indicates imidacloprid and the application 

methods do not affect tree growth. Treatment of high-risk host trees with 

imidacloprid potentially kills ALB. 

b. Environmental Resources 

(1) Water 

The removal of infested trees under this alternative has similar water 

quality impacts as Alternative 2—Removal of Infested Trees. Under this 

alternative, insecticide treatment would slow the spread of ALB, however, 

it will not stop the spread as treatment does not provide 100 percent 

protection to all trees. In comparison to Alternative 3—Full Host 

Removal, we expect this alternative to have less of a short-term impact to 

water quality in regards to host tree removal because the Program would 

remove fewer trees, at least initially, under this alternative. 

Alternative 4 would have the greatest potential for imidacloprid to move 

into surface or ground water because chemical treatment is the only option 

available for high-risk host trees. Imidacloprid exhibits physical and 

chemical properties that suggest it could contaminate surface and ground 

water. Detections in ground water have occurred in various parts of the 

United States, including States where ALB is present. Solubility and a 

lack of affinity for binding to soil or sediment suggest that imidacloprid 

could move offsite through runoff or leaching (appendix F). The ability to 

leach into ground water would depend on site-specific conditions, such as 

soil type and depth to the water table. However, label restrictions 

regarding applications near surface water and other information regarding 

ground water reduces the potential for water contamination. In addition, 

the preferred use of tree injections of imidacloprid by the Program further 

reduces the possibility of impacts to water quality when compared to soil 

injection. 

4. Alternative 

4: Infested 

Tree 

Removals 

and 

Insecticide 

Treatment of

High-Risk 

Host Trees 
76 IV. Environmental Consequences 

APHIS collected water samples as part of its monitoring effort to 

determine the potential for imidacloprid to move to surface and ground 

water from Program applications. (The data is available in appendix F.) 

The imidacloprid level in a majority of the water samples is below 

detection in surface and ground water, or below levels that would impact 

human health or the environment. However, detection of imidacloprid in 

some water samples is above the level the Program expects from their 

applications. Due to the widespread use of imidacloprid for other uses, 

especially home and garden, it is difficult to attribute detections to 

program applications in urban areas where a majority of program 

treatments have taken place. The State of New York, after finding 

imidacloprid in water samples from all uses, restricts the application of 

imidacloprid to trunk injection. 

(2) Air 

The impacts on air quality from the removal of infested trees (as described 

in Alternative 2—Removal of Infested Trees) are the same for this 

alternative. Treatment of ALB-host trees with imidacloprid usually occurs 

through trunk injection, however, sometimes through soil injection. The 

trunk injection method involves injecting imidacloprid directly into the 

tree trunk; this is unlikely to affect air quality due to minimal exposure of 

the insecticide to the ambient air. A minor amount of the insecticide 

volatilizes during the soil injection until full adsorption into the soil and 

uptake by the plant roots occurs. Drift is not expected to impact air 

quality because both application methods would not have any associated 

drift. The chemical and physical properties for imidacloprid suggest that 

there is a low probability of imidacloprid volatilizing into the atmosphere 

and impacting air quality (see appendix F). 

(3) Soil 

Physical impacts to soil from alternative 4 would be expected to be less 

than those described in alternative 3. Tree removal would not occur for 

high-risk host trees reducing the potential for soil disturbance that could 

occur during removal. There is the potential for impacts to soil as a result 

from imidacloprid treatments of high-risk host trees. Impacts to soil 

quality would be greatest for soil injections compared to tree injections. 

Sensitive soil terrestrial invertebrates would be impacted in the immediate 

area of treatment in the case of soil injection applications; however, soil 

injections rarely occur, and tree injections would minimize these types of 

impacts. Imidacloprid negatively affects earthworms in soil at 

concentrations that have been observed in previous ALB eradication 

efforts using soil injections; however, reported levels do not appear to 

cause long-term impacts to soil microbes (appendix F; (Tu, 1995; 

Kreutzweiser et al., 2008a)). 
IV. Environmental Consequences 77

c. Ecological Resources 

The impacts to ecological resources, under alternative 4, is expected to be 

less than those described under Alternative 1—No Action Alternative and 

Alternative 2—Removal of Infested Trees because of the prevention of 

ALB dispersal to non-infested trees. Impacts to ecological resources from 

fragmentation and loss of trees in riparian areas would be less than 

alternative 3 because ALB high-risk host trees would receive an insecticide 

treatment rather than being removed. Under this alternative, ecological 

impacts from tree removal would be reduced; however, there is the 

potential for impacts from the use of the insecticide, to these same 

resources that would not occur under alternatives 1–3. These impacts are 

summarized in the following section with a more detailed analysis in 

appendix F. 

There may be circumstances where imidacloprid use is not practical and, 

under this alternative, the inability to remove high-risk host trees would 

allow for the spread of ALB to other trees. In addition, imidacloprid 

efficacy is variable as a prophylactic treatment for ALB (Poland et al., 

2006). Site-specific conditions, regarding tree health and other factors, 

can impact the uptake and distribution of imidacloprid in trees, allowing 

ALB to survive even within imidacloprid-treated trees. In these situations, 

ALB would be able to spread, with the possibility of additional infested 

tree removal and insecticide treatments, increasing the risk to terrestrial 

and aquatic ecological resources. 

(1) Insecticide/Herbicide Use 

Imidacloprid has low to moderate acute toxicity to wild mammals, based 

on the available toxicity data. Imidacloprid is considered toxic to birds 

with acute oral median toxicity values ranging from 25 to 283 mg/kg 

(Anatra-Cordone and Durkin, 2005; EPA, 2014c). Reproduction studies 

using test species to support pesticide registration (e.g., the mallard and 

bobwhite quail) have shown NOECs of 125 ppm for both species. The 

application method and the available effects data indicate low exposure 

and risk to terrestrial vertebrates (e.g., birds and mammals) (appendix F). 

Applications using trunk and soil injection remove the risk of exposure 

from drift or runoff. There is the possibility of imidacloprid exposure to 

mammals and birds that may feed on insects or vegetation from treated 

trees. APHIS measured imidacloprid leaf and twig residue values, and 

demonstrated that most birds and mammals would have to consume 

several times their daily intake to reach an adverse effect. The risk to 

terrestrial vertebrates consuming insect prey with residues of imidacloprid 

is unknown; however, as these prey do not forage exclusively on treated 

trees and residue levels are not enough to kill them, the risk to predators is 

likely low. Imidacloprid is also specific to certain groups of insects, and 

would not affect all insects that are present on treated trees. 
78 IV. Environmental Consequences 

Technical and formulated imidacloprid is acutely toxic to honey bees, and 

other related bee species, at exposures of 3.7 to 230 nanograms (ng)/bee 

by oral and contact exposure (Schmuck et al., 2001; Tasei, 2002; AnatraCordone and Durkin, 2005; EPA, 2014a). Acute sublethal effects in 

laboratory studies show NOECs at less than 1 ng/bee (Anatra-Cordone and 

Durkin, 2005). Imidacloprid metabolite toxicity to honey bees is variable 

with some of the metabolites having equal toxicity to imidacloprid, while 

other metabolites are practically nontoxic (Anatra-Cordone and Durkin, 

2005). Several studies have been conducted to determine potential 

sublethal effects in laboratory and field situations. Studies to assess the 

effects of imidacloprid on homing behavior, colony development, foraging 

activity, reproduction, wax/comb production, and colony health, as well as 

other endpoints, generally reveal effects at levels above those measured in 

nectar and pollen in the field from agricultural crops under various 

application methods (appendix F). 

Impacts to honey bees from sublethal exposure to imidacloprid in the 

presence of other stressors have also been evaluated in laboratory studies. 

These studies suggest that pesticides, such as imidacloprid, in combination 

with pathogens may impact colony health and immune function in honey 

bees (Alaux et al., 2010; Vidau et al., 2011; Pettis et al., 2012). Due to the 

uncertainty of the risk to honey bees from the proposed treatments in this 

program, APHIS provided funding for a multiyear study to determine the 

potential for exposure and effects to honey bees from the proposed 

treatments (Johnson, 2012). Results from that work suggest that these 

types of applications do not adversely impact honey bees and their hives, 

and that imidacloprid residue pollen levels collected from maple trees is 

low, with an average concentration of 5.3 micrograms (μg)/kg from tree 

injection applications, and 0.28 μg/kg from soil injections of imidacloprid. 

Residues of imidacloprid and six associated metabolites were reported as 

below detection in nectar samples. 

Imidacloprid exposure to terrestrial invertebrates from the ALB 

Eradication Program, especially honey bees, is not expected to result in 

significant risk to pollinators. Pollinator exposure to imidacloprid is 

reduced because only treated trees and their associated flowers and pollen 

could have residues, while other flowering plants that have not been 

treated would not contain residues. Exposure and risk would increase in 

cases where large numbers of trees are treated over large areas prior to 

flowering, and in cases where only flowers from treated trees are the 

primary nectar source. This may occur in the Northeastern United States 

where maple trees bloom prior to many other flowering plants. 

Applications of imidacloprid, particularly via soil injection, could expose 

soil-dwelling invertebrates sensitive to the insecticide; however, 

environmental fate data indicate the effects are transient (Anatra-Cordone 

and Durkin, 2005). In cases where imidacloprid is tree-injected, there 
IV. Environmental Consequences 79

would be reduced exposure and risk to soil-dwelling terrestrial 

invertebrates; exposure would occur primarily from leaves that drop from 

treated trees. These risks would be proportional to the number of treated 

trees in a given area. 

Imidacloprid has low toxicity to aquatic organisms including fish, 

amphibians, and some aquatic invertebrates. Acute toxicity to fish and 

amphibians is low, with acute median lethal concentrations typically 

exceeding 100 mg/L (Anatra-Cordone and Durkin, 2005; EPA, 2014a). 

Chronic toxicity to fish is in the low ppm depending on the test species 

and endpoint. Aquatic invertebrates are more sensitive than fish to 

imidacloprid (Anatra-Cordone and Durkin, 2005; EPA, 2014a) (appendix 

F). 

Imidacloprid exposure in aquatic environments and risk to aquatic biota is 

low (appendix F). The method of application eliminates the potential for 

drift and, in the case of tree injections, eliminates the probability of offsite 

runoff. Another potential pathway of exposure to aquatic organisms is 

imidacloprid residues in leaf litter from treated trees. Aquatic 

invertebrates feeding on leaf litter containing imidicloprid residues have 

measurable sublethal impacts, as well as impacts on decomposition rates 

(Kreutzweiser et al., 2007; Kreutzweiser et al., 2008a, b; Kreutzweiser et 

al., 2009). Mortality to some leaf-shredding insects occurred when they 

were intentionally overdosed at levels higher than they would encounter at 

typical field applications (Kreutzweiser et al., 2007; Kreutzweiser et al., 

2009). Not all plant material available to aquatic decomposers will 

contain imidacloprid as ALB-host trees are part of a greater ecosystem 

that will contain other plant and organic material. Exposure and risk to 

aquatic organisms will increase in situations where large numbers of trees 

may be treated within a watershed. The risk to aquatic organisms from 

this exposure can be reduced by not treating trees, or treating a small 

number of trees, and avoiding treatments close to surface water. 

Bioconcentration or bioaccumulation of imidacloprid is not anticipated 

based on the chemical and environmental fate data and proposed use 

pattern (appendix F). 

There is a potential for subsurface transport of imidacloprid to aquatic 

habitats for applications made directly into soil. This exposure will be 

minimized by only making applications where the ground water table is 

not in proximity to the zone of injection, and avoiding soils that have a 

high-leaching potential. Conservative estimates of potential aquatic 

residues in static, shallow bodies of water from soil injections, based on 

maximum label rates, demonstrate values that are not expected to have 

indirect or direct impacts to aquatic biota (appendix F). Actual aquatic 

residues will be below levels that could harm aquatic biota due to the low 

chance of moving offsite, which is a result of the method of application 

and environmental fate of imidacloprid. 
80 IV. Environmental Consequences 

The potential impacts to ecological resources from herbicide use would be 

low and less than those described under alternatives 2 and 3 (appendix E) 

because fewer trees would be cut down. Expansion of infested areas or 

new infestations would occur at a slower rate under alternative 4, when 

compared to alternative 2. This would result in less herbicide use in an 

infested area, and in new introductions reducing the impacts to aquatic and 

terrestrial ecological resources. Herbicide use would be greater under 

alternative 3, compared to alternative 4, because more trees would be 

removed under alternative 3 (both infested trees and host trees within a 

1⁄2-mile radius). 

d. Economic, Social, and Cultural Resources 

The economic impacts, under alternative 4, would be less than those 

described under alternatives 1 and 2. In relation to preventing the spread 

of ALB, the economic impacts may be similar to those described under 

alternative 3 when imidacloprid treatments are successful in serving as a 

prophylactic treatment of high-risk host trees. However, due to the 

variability in imidacloprid uptake by trees and other site-specific 

conditions (e.g., general tree health), insecticide treatments may not 

provide the level of protection that removal of high-risk host trees would 

provide. In cases where imidacloprid treatments cannot be successfully 

used for high-risk host trees, the economic impacts would be greater than 

those under alternative 3 because ALB would be able to expand its range 

and infest new areas. In addition, the cost of multiple insecticide 

treatments to a tree is greater than the costs of removal. 

Potential economic losses specific to the use of imidacloprid could occur 

for the maple syrup industry because imidacloprid label restrictions would 

render treated trees unusable for syrup production. Imidacloprid is a 

systemic insecticide, and concentrations of the insecticide will be present 

in the tree sap. The economic impact to maple syrup producers will be 

dependent upon the proportion of their maple trees removed from 

production through either tree removal or treatment with imidacloprid. 

Similarly, nursery tree producers would incur costs by treating host trees 

with imidacloprid to meet regulatory requirements for shipping trees 

outside the quarantine boundary. Sales of nursery trees that are hosts to 

ALB, regardless of treatment with imidacloprid, may drop as buyers select 

trees that are not hosts to ALB. 

As discussed in alternative 2, the removal of infested trees from public and 

private lands may result in economic, social, and cultural impacts. The 

option to be able to treat trees with imidacloprid will be based on an 

evaluation, by the Program, of site-specific conditions for a given 

infestation. 
IV. Environmental Consequences 81

The forestry and tourism industries, as well as public and private 

landowners located near an ALB quarantine area, will gain protection 

from ALB through the confinement of pest populations, and eradication 

through removal of infested host trees and treatment of host trees in 

proximity to infested trees. 

e. Human Health 

Under alternative 4, in the short term, tree loss would occur from removal 

of infested host trees. Initially, the number of trees lost from removal 

activities is similar to alternative 2 as only infested host trees would be 

removed in this alternative. Prophylactic treatments with imidicloprid 

may protect some high-risk host trees and may result in an overall 

decrease in the removal of infested trees compared to alternative 2. 

APHIS expects the removal of fewer trees under alternative 4 compared to 

alternative 3—Full Host Removal. Some localized negative human health 

consequences from tree loss are expected as discussed in alternative 2. 

This alternative is less effective than alternative 3 because insecticide 

treatment is unlikely to stop the spread of ALB resulting in additional tree 

loss in the future.

(1) Insecticide/Herbicide Use 

APHIS uses the insecticide imidacloprid through trunk or soil injections to 

protect trees from ALB infestation. For trunk injections, applicators drill 

holes around the trunk, 2 to 6 inches above the soil-wood line. For soil 

injection, applicators inject imidacloprid, at a minimum of four injection 

sites spaced evenly around the base of the tree. Application occurs under 

the soil around the base of the tree, normally no more than 12 inches from 

the base. In addition, APHIS uses herbicides to treat stumps and roots of 

felled trees, as described in alternative 2. 

APHIS prepared a risk assessment to evaluate the risks for workers and 

the general public associated with the use of imidacloprid applications 

using trunk and soil injections (appendix F). Several formulations are 

available for this particular application. Technical and formulated 

imidacloprid (Merit® 2F) has low to moderate acute oral mammalian 

toxicity, with median toxicity values ranging from 400 to greater than 

4,000 mg/kg. Acute lethal median toxicity values are typically greater 

than 2,000 mg/kg and 5.3 mg/L for dermal and inhalation exposures, 

respectively. Chronic oral exposure toxicity studies in rat, dog, and mouse 

showed that the rat was the most sensitive test species with a NOEL of 

5.7 mg/kg/day. The EPA acute and chronic reference doses (RfD) are 

0.14 mg/kg/day and 0.057 mg/kd/day, respectively. 

Imidacloprid is a neurotoxic insecticide, based on its mode of action. 

Available literature for imidacloprid and associated metabolites suggest a 
82 IV. Environmental Consequences 

lack of carcinogenic, mutagenic, or genotoxic effects at relevant doses. 

Developmental, immune, and endocrine-related effects were observed in 

some mammal studies; however, these effects occurred at doses that 

would not be expected to occur under program use. 

Exposure to imidacloprid is greatest for applicators, but is reduced by 

following label directions regarding personal protective equipment (PPE). 

The human health risk assessment quantified risks for occupational worker 

exposure and general public exposure using a child soil ingestion (pica) 

scenario. Results from that assessment show that imidacloprid used to 

control ALB poses minimal risk to human health under both exposure 

scenarios. The human health risk assessment (included in appendix F) 

provides details on toxicity, exposure, and risk associated with 

imidacloprid. 

Herbicide risks to human health would be similar to those described under 

alternative 2. Notification to landowners also occurs in the case of any 

chemical treatments that may be used to kill stumps (herbicides) or treat 

high-risk host trees for ALB (insecticide). In addition to the notification 

process, there are other risk reduction requirements for pesticide use that 

are designed to protect workers, as well as the general public and the 

environment. 

Another potential exposure pathway for the public is the use of 

imidacloprid-treated trees for firewood. The levels of imidacloprid in 

treated trees that could be used as firewood is expected to be low because 

the insecticide moves to the leaves, and smaller, actively growing 

branches in the tree where insect feeding is greatest; these parts of the tree 

would not typically be used as firewood. In cases where trees are treated, 

their removal would not be expected to occur in the same growing season 

as treatment, allowing degradation of imidacloprid. In addition, trees 

harvested for firewood are usually allowed to dry before they are used as 

fuel, which would allow for additional degradation of imidacloprid. The 

rapid combustion of wood at high temperatures, as can occur in a fireplace, 

results in rapid degradation of other types of pesticides; residues are more 

likely under slow combustion and temperatures less than 1112 o

F (600 oC) 

(McMahon et al., 1985; Bush et al., 1987a; Bush et al., 1987b). 

Imidacloprid would be expected to degrade at temperatures similar to 

those that would occur from burning firewood, based on its measured 

thermal decomposition temperature, which is below 932 oF (500 oC). 

Potential thermal degradation products from the use of the imidacloprid 

formulations that could be used in the Program include hydrogen cyanide, 

carbon monoxide, and oxides of nitrogen and carbon. Concentrations of 

these degradation products would be very low due to the expected 

concentrations of imidacloprid in firewood, and potential temperatures 

that could occur in burning firewood. 
IV. Environmental Consequences 83

a. Forest Resources 

The impacts of the preferred alternative on forestry resources will be 

similar to those impacts described in alternatives 3 and 4. Under the 

preferred alternative, the Program may either remove high-risk host trees 

or treat high-risk host trees with imidacloprid. An integrated approach 

alternative allows for the adaptation of eradication methods to suit sitespecific needs and resources. In the long term, adverse impacts to forest 

resources would be greatest under Alternative 1—No action Alternative 

and Alternative 2—Removal of Infested Trees when compared to the 

preferred alternative, the integrated approach. Alternative 1 is not an 

eradication program and ALB would continue to spread under 

alternative 2. 

b. Environmental Resources 

(1) Water 

Impacts to water quality (described for the previous alternatives, other 

than the no action alternative), would apply to the preferred alternative as 

they require some level of tree removal that can result in impacts to water 

quality. These impacts will be greatest in watersheds that have soils that 

are vulnerable to erosion and large numbers of host trees in proximity to 

water. However, the short-term impacts to water quality would be 

reduced under the preferred alternative when compared to the 

implementation of the exclusive use of tree removal or insecticide 

treatment of high-risk host trees. 

Flexibility in addressing high-risk host trees would allow the Program to 

implement measures with the greatest chance of success based upon sitespecific conditions and resources. Potential long-term impacts to water 

quality would be reduced as ALB eradication efforts would have a greater 

probability of success, and prevent the spread and introduction of ALB to 

other watersheds in the United States where host trees are present. 

(2) Air 

The impacts to air quality (described for the previous alternatives, other 

than the no action alternative) would apply to the preferred alternative, as 

it requires some level of tree removal, which can affect air quality. 

However, the impacts for the preferred alternative would be reduced 

compared to those alternatives because neither imidacloprid treatment nor 

removal of high-risk hosts would be used exclusively. Impacts to air 

quality from tree removal activities would be more localized when 

compared to full host tree removal. Air quality impacts from the use of 

the herbicides, or the insecticide imidacloprid, are not anticipated based on 

5. Alternative 

5: Integrated 

Approach 

(Preferred 

Alternative) 
84 IV. Environmental Consequences 

their use pattern and the low probability of any chemical volatilizing into 

the atmosphere. 

(3) Soil 

The impacts to soil would be similar to those described for alternatives 3 

and 4, Full Host Removal and Insecticide Treatment of High-Risk Host 

Trees. Both Alternative 1—No Action and 2—Removal of Infested Trees 

would allow ALB to continue to expand its distribution within the United 

States, with impacts to soil quality dependent upon the importance of 

ALB-host species in soil nutrient cycling. The potential for soil quality 

impacts would increase with the removal of infested trees under 

alternatives 2 through 4; however, impacts under the preferred alternative 

would be less as there would be an option for tree removal of ALB highrisk host trees or insecticide treatment. Insecticide treatment can impact 

soil quality for those soil fauna that are sensitive to imidacloprid; 

nonetheless, these impacts would be more localized than tree removal 

activities that can result in the physical disruption of soil through the use 

of heavy equipment. 

c. Ecological Resources 

Environmental impacts from the selection of the preferred alternative may 

result in long-term impacts that are less than those described under the no 

action alternative and the other eradication alternatives. The 

implementation of alternative 5 has been shown to be a successful 

eradication program strategy in site-specific infestations. Risks to 

ecological resources would be localized to the areas of infestation similar 

to alternatives 3 and 4; however, in the case of high-risk host trees, the 

Program would have the flexibility to select the treatment option that best 

fits site-specific conditions. There would be potential risks related to 

removal of trees, as described under alternative 3, and insecticide 

treatment, described under alternative 4 however, flexibility to select 

either method would reduce the risk when compared to the exclusive use 

of either eradication method.

d. Economic, Social, and Cultural Resources 

Selection of the preferred alternative may have economic, social, and 

cultural impacts similar to those described for alternatives 3 and 4. These 

impacts would be reduced under the preferred alternative because there is 

flexibility in how high-risk host trees would be treated. Alternative 1 and 

alternative 2 would have the greatest potential impacts because neither 

approach eradicates ALB. 
IV. Environmental Consequences 85

e. Human Health 

Under alternative 5, the rate of tree loss is expected to be reduced the most 

compared to the other alternatives. In the short term, tree loss may occur 

from removal of infested and some surrounding host trees. The amount of 

tree loss from the removal activities would be more than alternative 2 and 

4 but less than alternative 3. Localized negative human health 

consequences from tree loss (e.g., increased stress and reduced air quality 

as discussed in the no action alternative) are expected in the affected areas 

where ALB infested and high-risk host trees are removed. 

This alternative is expected to eradicate ALB resulting in less tree loss in 

the long term. Eradication of ALB would allow all the benefits of trees to 

continue in broader areas, and the negative consequences from a lack of 

trees can be avoided. Specifically, shading and improved air quality 

would continue to be available for people living in and travelling through 

affected areas. Trees would continue to provide opportunities for stress 

reduction and beautify the landscape. 

f. Adaptive Management Approach 

The adaptive management approach evaluates future pesticide use for 

similar or less impacts than those described for the pesticides in each of 

the alternatives discussed in this EIS. This includes any changes in 

herbicide or insecticide use that could occur in the future as the Program 

evaluates new chemical treatment options. The use of the adaptive 

management approach is only in cases where a human health and 

ecological risk assessment demonstrates that proposed pesticides would 

have equal or less risk to human health and the environment, in 

comparison to the pesticides currently used by the program. The criteria 

for this approach ensures that the potential pesticide-related impacts to the 

various resources discussed in this EIS are similar, or less, than those 

associated with any new pesticide use proposed in the future. 

Cumulative impacts are those impacts on the environment that result from 

the incremental impact of an action when added to other past, present, and 

reasonably foreseeable future actions, regardless of what agency (Federal 

or non-Federal) or person undertakes such actions. Cumulative impacts 

can result from individually minor, but collectively significant actions 

taking place over time. The potential for cumulative impacts in the ALB 

Eradication Program will vary across the United States due to the various 

site-specific conditions where an infestation may occur. 

The cumulative impacts discussion in this EIS is meant to be general 

because the baseline environmental conditions vary between urban and 

forested areas, and other site-specific conditions that may occur where 

ALB-host trees are present. Site-specific EAs may incorporate this 

6. Cumulative 

Impacts 
86 IV. Environmental Consequences 

information to better characterize the potential cumulative impacts related 

to an ALB Eradication Program. 

For the Program, the cumulative impacts are primarily associated with the 

loss of trees, and the proposed use of an insecticide and herbicides. The 

severity and intensity of tree loss varies under the different alternatives, as 

well as the dominance of ALB-host trees in a given area. 

Cumulative impacts related to forest resources would be expected under 

those alternatives where ALB is likely to spread and infest new areas. 

Cumulative impacts to forest resources would also be expected during 

successful eradication efforts; however, those impacts would be more 

localized and short term compared to instances where ALB is allowed to 

spread. The effects of natural and manmade stressors to forests (e.g., 

timber harvests, acid rain, climate change, and other pests and diseases) 

can be additive or synergistic, that is, the effects of all of the stressors 

together become greater than the individual stressors alone (Cox, 1999; 

Logan et al., 2003). The effect of these other stressors added to the impact 

of an ALB infestation could increase the severity of an ALB outbreak. An 

example of this type of cumulative impact is evident with the decline of 

oaks in the Ozark forests in the Central United States. A native 

cerambycid beetle, the red oak borer (Enaphaledes rufulus) was not a 

major pest of oak, but is now considered a major contributor to oak 

decline in the area due to other factors that predispose the trees to 

infestations at higher levels than previously recorded (Coulson and 

Stephen, 2006). 

The loss of ALB-host trees may have cumulative impacts by increasing 

the ability of invasive plants to establish in areas where canopy gaps 

occur, or other areas where host trees are lost. These types of 

introductions have been noted in disturbed forested areas, and may impact 

natural plant succession, as well as nutrient cycling (Woods, 1997; 

Meiners et al., 2002). 

Currently, FS is addressing approximately 70 invasive species (plants and 

insects) in areas where ALB-host trees are present. The impact from the 

loss of ALB-host trees could result in cumulative impacts that would 

benefit the establishment of many of these species. These impacts would 

be more widespread under the no action alternative (1) and alternative 2 

because ALB would be able to establish and spread to other areas of the 

United States. These types of cumulative impacts may also occur under 

alternatives 3 through 5 because tree removal would occur, however, for 

the alternatives that result in eradication, the potential cumulative impacts 

would be reduced. 

Under the no action alternative, it is possible other State and local agencies 

or tribes will implement an eradication program without involvement from 
IV. Environmental Consequences 87

APHIS. Given the limited jurisdiction of these entities, the eradication 

programs would lack regional or national coordination, which could result 

in the spread of ALB to other host trees over a larger geographic area. 

The expansion of ALB beyond the current areas could result in additional 

stressors to host trees, causing both economic and environmental impacts. 

Abiotic and biotic stressors (e.g., climate change, other invasive pests, and 

air pollution) all pose threats to ALB-host trees; the addition of ALB to 

urban and natural forest ecosystems would be expected to result in 

cumulative impacts beyond those already identified as potential stressors 

(Horsley et al., 2002; Poland and McCullough, 2006; Iverson et al., 2008). 

Economic losses to the timber and maple syrup industries would be 

anticipated, as well as increased costs to homeowners that choose to treat 

trees or have them removed once they are infested. Economic data for the 

loss of ash trees in Ohio from an invasive pest, which is just one of the 

hosts for ALB, show that landscape loss, tree removal, and replacement 

costs could range between $1.8 and $7.6 billion due to the emerald ash 

borer (Agrilus planipennis) (Sydnor et al., 2007). The loss of one tree 

species is significant; ALB has many more host trees than emerald ash 

borer. 

Cumulative impacts to the environment would also be expected as ALBhost trees are lost from urban and natural forests. The potential for 

cumulative impacts would be dependent upon whether trees are removed 

from urban or forested areas, and their dominance in those areas. Tree 

removal, under the various alternatives, could result in impacts to soil and 

water quality. These impacts are reduced by implementation of BMPs. In 

urban areas, residential and business development result in soil 

disturbance and potential water quality impacts due to land disturbance, 

and increase impervious surfaces which have the ability to transport a 

variety of pollutants to surface and ground water. 

ALB eradication efforts in urban areas are expected to have incrementally 

minor impacts to environmental quality when put in the context of other 

activities that may impact air, soil, and water. In forested areas, the 

cumulative impacts from tree removal may result in cumulative impacts to 

soil and water quality, resulting in increased erosion and transport of 

sediments and nutrients to water and, in particular, in cases where large 

areas of timber are removed in proximity to water, or from watersheds 

vulnerable to soil erosion. The magnitude of these contributions to natural 

and other man-made sources of sediment and nutrients will vary, 

depending on site-specific conditions and other activities that may be 

occurring in a given watershed. The potential for cumulative impacts in 

these scenarios would be reduced due to the implementation of BMPs by 

the Program, where applicable. The selection of an alternative that will 

ensure eradication will reduce the likelihood of cumulative impacts 

occurring over a larger geographic area. There would be some loss of 

wildlife habitat in areas where host trees are removed; however, those 
88 IV. Environmental Consequences 

losses would not be considered permanent because in unmanaged habitats 

(e.g., woodlots), stumps of high-risk host trees would be allowed to 

resprout, and replanting activities may occur in managed areas. 

Cumulative impacts to human health and the environment are anticipated 

to be incrementally minor for the proposed use of herbicides. All three 

herbicides have agricultural and non-agricultural (e.g., right-of-way and 

forestry) uses. FS uses triclopyr and, to a lesser extent, imazapyr in many 

of its invasive weed control programs where ALB-host trees are present. 

The proposed use of herbicides in the ALB Eradication Program is not 

expected to contribute significantly to the overall use of herbicides by 

other entities. The Program applies herbicides to stumps or sprouting 

vegetation from cut stumps, using hand painting or backpack spray 

applications that minimize offsite transport of the proposed formulations. 

Recent ALB eradication efforts suggest that stump removal is the 

preferred treatment method, with greater than 75 percent of the stumps 

being removed rather than being treated with an herbicide. 

 

Imidacloprid is widely used in urban and agricultural settings; however, 

the increase in loading beyond current use, in addition to that which could 

be added due to ALB treatments, is difficult to quantify because the 

number of treated trees is unknown relative to current use patterns for 

areas where ALB-host trees may occur. In addition, the use of 

imidacloprid under the preferred alternative would vary because of the 

choice of tree removal and insecticide treatment. 

Cumulative impacts from the proposed use of imidacloprid would not be 

expected to result in significant impacts to human health. The toxicity 

profile, method of application, and notification to the public regarding its 

use would minimize exposure and risk to the public. Adherence to label 

requirements would also minimize exposure and protect human health for 

workers involved with ALB eradication efforts. 

The amount of imidacloprid added to the environment would be greatest 

under alternative 4 because high-risk host trees within a 1⁄2-mile radius of 

infested trees would receive treatment, compared to the preferred 

alternative where only select trees would receive imidacloprid treatments. 

The cumulative risk to aquatic resources would be greatest when 

considering large-scale imidacloprid treatments of deciduous trees, such as 

ALB-host trees. Imidacloprid residues in leaf litter from treated trees can 

be transported to aquatic environments and result in sublethal impacts to 

some aquatic invertebrates (Kreutzweiser et al., 2007; Kreutzweiser et al., 

2008b; Kreutzweiser et al., 2009). In those studies, the more significant 

impacts occurred in cases where exposure to imidacloprid in leaf litter was 

at concentrations greater than anticipated under the current proposed use 

pattern. These impacts are selective to certain types of aquatic 
IV. Environmental Consequences 89

invertebrates due to their feeding preference, and would not be anticipated 

for other aquatic invertebrates. Cumulative impacts from the addition of 

treated leaves would be expected to only occur for some aquatic 

invertebrates in cases where host trees are a dominant species and large 

numbers are treated. The cumulative impacts to aquatic invertebrates 

would also be reduced by the presence of non-treated leaf litter and other 

organic matter present in aquatic habitats. 

Streams that may already be impacted by other factors could have 

cumulative impacts related to imidacloprid use in cases of large-scale 

treatments. Available water monitoring data for imidacloprid and other 

neonicotinoid insecticides in the United States shows that detections are 

common in urban and agricultural areas (Phillips and Bode, 2004; Starner 

and Goh, 2012; Hladik et al., 2014). These detections occur along with 

multiple other pesticides, and other organic and inorganic contaminants. 

Water quality data show pesticide mixtures to be a common occurrence. 

Water quality data collected by the U.S. Geological Survey, as part of a 

national monitoring effort, shows that 50 percent of the surface water 

samples contain four or more pesticides in urban, agricultural, and mixeduse watersheds (Gilliom et al., 2007). The impact of these mixtures on 

human health and aquatic ecosystems is poorly understood. Chemical 

mixtures may have additive, synergistic, antagonistic, or potentiation 

effects on biological systems. Studies testing imidacloprid and other 

chemical mixtures have reported results showing no interactions between 

chemicals or additive, synergistic, and antagonistic effects to test 

organisms (Chen et al., 2010; Loureiro et al., 2010; Zhang et al., 2010; 

Pavlaki et al., 2011). The chemical mixture, timing and level of dosing, 

and endpoints measured were all variables that can impact how these 

chemicals may act in combination. Interactions between pesticides, such 

as imidacloprid and other stressors (e.g., water quality and predators) have 

also been reported for various aquatic and terrestrial species (Holmstrup 

et al., 2010; Laskowski et al., 2010). 

Uncertainties regarding the prediction of effects to nontarget organisms 

from imidacloprid and all possible mixtures, as well as the large area of 

the United States that is being considered in this EIS, make it difficult to 

determine cumulative impacts from program insecticide treatments. 

However, the incremental increase in risk due to imidacloprid use is 

expected to be minor for aquatic communities because the potential for 

imidacloprid risk to aquatic habitats is low; this is based on the proposed 

method of application, environmental fate, and available information 

regarding imidacloprid effects to aquatic organisms. 

Cumulative impacts to terrestrial wildlife associated with imidacloprid use 

are also expected to be incrementally minor for most animals. 

Imidacloprid residues on food items for mammals and birds are not 
90 IV. Environmental Consequences 

expected to result in significant risk to wildlife. The low risk to mammals 

and birds suggests that any cumulative impacts of imidacloprid would be 

minor when put in context with other stressors, such as loss of habitat. 

Imidacloprid will affect sensitive nontarget invertebrates that consume 

plant material from treated trees; however, the mode of action of 

imidacloprid, and the method of application, which is targeted to certain 

trees, will reduce impacts to most nontarget invertebrates. The risk to 

sensitive terrestrial invertebrates from proposed imidacloprid applications 

is expected to be minor in relation to other stressors, such as other 

pesticide applications and loss of habitat. The lack of significant 

cumulative impacts to most terrestrial invertebrates will also ensure that 

mammals and birds that prey on insects are not impacted. 

Large-scale treatment of trees using imidacloprid could also increase 

pesticide exposure to pollinators above current levels. A variety of factors 

stress native pollinators, as well as honey bees. Stressors include 

environmental pollution, habitat loss, poor nutrition, pests and diseases 

(e.g., Varroa mites on honey bees), and some pesticides, including 

imidacloprid (Potts et al., 2010; USDA, 2012; Sanchez-Bayo and Goka, 

2014). Recent studies have shown that honey bees exposed to sublethal 

concentrations of insecticides and pathogens indicate interactive negative 

effects from exposure (Alaux et al., 2010; Pettis et al., 2012). 

Interactions between imidacloprid, as well as other neonicotinoids and 

pathogens (e.g., Nosema), have resulted in colony and immune function 

impacts to honey bees (Alaux et al., 2010; Vidau et al., 2011; Pettis et al., 

2012). Colony collapse disorder (CCD) is a problem impacting domestic 

honey bee populations, with potentially significant impacts regarding 

pollination of agricultural and native plants. The causal agents for CCD 

are a complex variety of stressors, of which imidacloprid and other 

insecticides may contribute (USDA, 2012; Lu et al., 2014). The potential 

for exposure and cumulative impacts to honey bees, and other pollinators, 

from imidacloprid use is reduced by the availability of other species of 

flowering plants, and selectively treating trees in the ALB quarantine area. 

B. Special Programmatic Considerations 

a. Endangered Species Act 

Section 7 of the ESA and its implementing regulations require Federal 

agencies to ensure their actions are not likely to jeopardize the continued 

existence of listed threatened or endangered species, or result in the 

destruction or adverse modification of critical habitat. 

 

1. Applicable 

Environmental 

Statutes 
IV. Environmental Consequences 91

(1) Potential Effects of ALB Establishment on Listed Species

ALB has the potential to affect listed species and their habitats throughout 

the United States. ALB-host species, including maple, poplar, and birch, 

among others, play a critical role in the life histories of numerous listed 

species; impacts to these host species could affect their survival and 

recovery. Approximately 60 federally listed species and their critical 

habitats could be impacted by the introduction and spread of ALB in the 

United States based on the co-occurrence of preferred ALB-host trees and 

forest-dependent listed species in the Northeastern and mid-Atlantic areas 

(figure 4–2). Approximately half of the species in these areas are listed 

freshwater mussels, while approximately 18 percent are listed terrestrial 

plants. The remaining listed species are primarily mammals and fish. 

Widespread establishment of ALB in the United States will cause 

significant ecological damage. Impacts to terrestrial and aquatic fauna and 

flora are expected, especially in areas where ALB-host plants are 

prevalent. T&E species that depend on ALB-host trees would be most 

affected. The extent of damage and its impact on ecological community 

function would be dictated by multiple factors. In cases where the host 

trees are not a dominant component of the forest stand, impacts may be 

less; however, in situations where host trees occur in high density and are 

keystone species within that particular forest type, the impacts are almost 

certain to be significant. 

(2) Potential Effects of the ALB Program on Listed Species

Effects to listed species from the proposed alternatives for eradicating 

ALB may also pose a risk to protected species and their designated critical 

habitat without proper mitigation. A recent example of these impacts is 

with listed bat species that may occur where ALB eradication efforts are 

currently implemented. Bat species in the program area (e.g., Indiana bat 

and northern long-eared bat) use trees for roosting, as travel corridors, and 

as foraging habitat. A portion of these trees could be ALB-host trees that 

may require removal. 

Direct effects are those effects that are the result of the direct or immediate 

effects of the project on the species or its habitat. For bats, these effects 

can include increases in disturbance (i.e., in the form of noise, human 

activity, and vibrations from equipment) from tree clearing and would 

vary in intensity, depending upon the source (chainsaws, chippers, stump 

grinders, trucks). Tree removal can result in loss of maternity and nonmaternity roost trees. Removal of roost trees could impact bats by 

requiring them to expend time and energy to identify an alternate roost 

site. Removal of an unidentified roost tree during the summer occupancy 

season could result in crushing or injury of bats. 

 
92 IV. Environmental Consequences 

Figure 4–2. Federally listed threatened and endangered species by county that co-occur with ALB-host 

trees in the Northeastern United States. 

Indirect effects to bat species are those effects that are caused by or would 

result from the proposed action and are later in time, but are still 

reasonably certain to occur. These effects could result from reduction of 

bat insect prey from exposure to insecticides or herbicides. Tree removal 

may also indirectly affect bats by decreasing roost and habitat availability, 

and requiring additional time to forage and search for suitable roost trees. 

Elimination or significant reduction of fence rows and tree lines (used as 

travel corridors) may require bats to expend additional energy to find 

alternative routes. 

The removal of ALB-host trees can also impart beneficial effects to bats 

by creating canopy openings and more open understory, providing a 

higher quality habitat and an increase in maternity trees. Beneficial effects 

to bats from removal of ALB-host trees could create canopy openings in 

the interior of forests; this would increase solar exposure that may allow 

these larger trees to become suitable as maternity habitat. In addition, 
IV. Environmental Consequences 93

many of the trees removed are smaller trees; this would create a more open 

understory, providing higher quality habitat. 

Other listed terrestrial species may also be impacted by ALB eradication 

activities. Physical disturbance of areas where ground-nesting listed 

species occur can also have direct and indirect impacts. For example, 

disturbance of nesting areas of listed birds during the breeding season 

could occur from program activities near breeding areas. Disturbance can 

result in temporary abandonment of nests, exposing adults to aerial 

predation, and eggs and chicks to predation and inclement conditions. 

Tree removal activities may also result in take of listed species that may 

not be able to disperse during tree removal activities, and could be crushed 

by equipment. Forest-inhabiting plant species (e.g., running buffalo 

clover and small whorled pogonia) may be affected by clearing infested 

trees that would decrease forest floor shade, thus reducing habitat 

suitability. Tree removal could also result in trampling or physical 

destruction of plant populations. Insecticides applied to host trees could 

reduce pollinators of listed plant species. Program application of 

herbicides to stumps could directly harm listed plants if they are near 

treatments. 

ALB-host trees serve to protect water quality; their removal can 

destabilize stream banks resulting in excess sedimentation which would be 

detrimental to listed freshwater mussels (e.g., fanshell and pink mucket 

pearlymussel), fish, or other aquatic species in the vicinity of an ALBinfested area. As previously mentioned, over half of the listed species in 

the Northeastern United States that could be impacted by the loss of 

ALB-host trees are aquatic. Sedimentation and changes to water quality 

from tree loss (temperature, shading, and pesticide residues) may result in 

impacts to aquatic species due to program activities. 

(3) Consultation History

APHIS considers whether listed species, species proposed for listing, or 

critical habitat are present in the proposed program area. If none are 

present, no Section 7 consultation is required. If species or critical 

habitat is present in the proposed program area, APHIS conducts 

Section 7 consultation with the FWS and NMFS, on a site-specific 

basis, for ALB eradication activities. 

For the ALB Eradication Program in Worcester County, Massachusetts, 

APHIS consulted informally with FWS on a threatened plant, the small 

whorled pogonia, Isotria medeoloides, in 2008. In 2011, APHIS consulted 

informally with FWS on the impact of the eradication program on the 

small whorled pogonia and piping plover (Charadrius melodus) in Norfolk 

and Suffolk Counties in Massachusetts. 
94 IV. Environmental Consequences 

In June, 2011, APHIS first contacted the FWS in Columbus, Ohio for 

technical assistance regarding impacts to federally listed species in 

Clermont County, Ohio. Seven endangered species (Indiana bat, Myotis 

sodalis; running buffalo clover, Trifolium stoloniferum; fanshell, 

Cyprogenia stegaria; rayed bean, Villosa fabalis; snuffbox, Epioblasma 

triquetra; pink mucket pearlymussel, Lampsilis abrupta; and sheepnose, 

Plethobasus cyphyus) occur in Clermont County. Since that time, APHIS 

submitted biological assessments and consulted with FWS on those 

species, receiving concurrence with their “may affect, not likely to 

adversely affect” determinations, with the implementation of protection 

measures. In addition, FWS personnel made site visits to the infested area 

and have provided Indiana bat training to APHIS and the Ohio 

Department of Agriculture personnel. Also, APHIS conducted surveys for 

the Indiana bat, and other bat species, in the Clermont County eradication 

work zones and reported those findings to FWS. Most recently, APHIS 

entered into a formal consultation with FWS on the Indiana bat (receiving 

a biological opinion, dated June 4, 2014) and a conference on the northern 

long-eared bat (Myotis septentrionalis), a species proposed for listing as 

endangered. APHIS received a northern long-eared bat conference 

opinion on July 3, 2014. 

APHIS would continue to consult with FWS or NMFS, as necessary, 

when a known ALB infestation has been confirmed. In addition, APHIS 

would implement measures prior to the initiation of program activities to 

protect federally listed species and critical habitat. 

b. Migratory Bird Treaty Act 

The Migratory Bird Treaty Act of 1918 (16 U.S.C. 703–712) established a 

Federal prohibition, unless permitted by regulations, to pursue, hunt, take, 

capture, kill, attempt to take, capture or kill, possess, offer for sale, sell, 

offer to purchase, purchase, deliver for shipment, ship, cause to be 

shipped, deliver for transportation, transport, cause to be transported, 

carry, or cause to be carried by any means whatever, receive for shipment, 

transportation or carriage, or export, at any time, or in any manner, any 

migratory bird or any part, nest, or egg of any such bird. 

Prior ALB eradication efforts and consultations with FWS have resulted in 

several management recommendations to minimize impacts to migratory 

birds. These include: 

 minimize tree removals during nesting season 

 

 minimize disturbance as much as possible (avoid impacts to areas of 

non-host shrub/brush areas) 

 

 replant areas that have been significantly deforested 
IV. Environmental Consequences 95

 use existing trails for equipment to avoid disturbance to pastures/open 

fields that could be used as breeding sites for ground-nesting birds 

 have the names and contact information for local wildlife rehabilitators 

so that if there is an issue (e.g., as a raptor nest or fledging in the area), 

guidance can be provided regarding how to handle the situation 

c. Bald and Golden Eagle Protection Act 

The Bald and Golden Eagle Protection Act (16 U.S.C. 668–668c) prohibits 

anyone, without a permit issued by the Secretary of the Interior, from 

“taking” bald eagles, including their parts, nests, or eggs. The Act 

provides criminal penalties for persons who “take, possess, sell, purchase, 

barter, offer to sell, purchase or barter, transport, export or import, at any 

time or any manner, any bald eagle...[or any golden eagle], alive or dead, 

or any part, nest, or egg thereof.” The Act defines “take” as “pursue, 

shoot, shoot at, poison, wound, kill, capture, trap, collect, molest or 

disturb.” 

Without the implementation of the protective measures outlined below, 

tree cutting could disturb nesting eagles. FWS has recommended buffer 

zones from active nests which require different levels of protection (FWS, 

2007). They are as follows: 

 Avoid clear-cutting or removal of over-story trees within 330 feet 

of a nest at any time. (The Program will not use clear-cutting 

under any alternative discussed in this document.) 

 Avoid timber harvest operations (including road construction, and 

chain saw and yarding operations) during the breeding season, 

within 660 feet of the nest. The distance may be decreased to 330 

feet around alternate nests within a particular territory— 

 including nests that were attended during the current 

breeding season but not used to raise young, and 

 after eggs are laid in another nest within the territory have hatched. 

To ensure that program activities do not disturb eagles, APHIS would 

contact State fish and wildlife agencies, as well as the FWS, to determine 

the location of eagle nests in the program area. APHIS would also work 

with the FWS prior to any tree removal during the breeding season within 

660 feet of a nest to confirm that all eagles have left the nest. Outside of 

the breeding season, cutting may occur within the buffer zone around 

nests. 
96 IV. Environmental Consequences 

f. National Historic Preservation Act 

Section 106 of the National Historic Preservation Act requires Federal 

agencies to evaluate the potential impacts of their actions on historic 

properties. Historic properties are those that are on the National Register 

of Historic Places or meet the criteria for the National Register (NHPA, 

2014). To date, the ALB Eradication Program has met its Section 106 

responsibilities by contacting the State Historic Preservation Office 

(SHPO) in each State where eradication activities may occur. Due to the 

programmatic approach used in this EIS, and the uncertainty regarding 

where an ALB infestation may occur, APHIS has not contacted SHPOs in 

States where infestations have not occurred. APHIS will continue its 

current policy of evaluating program impacts to historical properties, and 

working with the SHPO in States with known ALB infestations. 

g. Executive Order 13045 

Executive Order (EO) 13045, “Protection of Children from Environmental 

Health Risks and Safety Risks,” acknowledges that children, as compared 

to adults, may suffer disproportionately from environmental health and 

safety risks due to their developmental stage, greater metabolic activity 

levels, and their behavior patterns. This EO (to the extent permitted by 

law and consistent with the agency’s mission) requires each Federal 

agency to identify, assess, and address environmental health risks and 

safety risks that may disproportionately affect children. An analysis of the 

potential impacts to children from the proposed alternatives, including the 

available human health risk assessments for proposed program pesticide 

use, suggest that no disproportionate risks to children are anticipated. 

h. Executive Order 12898 

EO 12898, “Federal Actions to Address Environmental Justice in Minority 

Populations and Low-Income Populations,” focuses Federal attention on 

the environmental and human health conditions of minority and lowincome communities, and promotes community access to public 

information and public participation in matters relating to human health 

and the environment. This EO requires Federal agencies to conduct their 

programs, policies, and activities that substantially affect human health or 

the environment in a way so as not to exclude persons and populations 

from participation in or benefiting from such programs. It also enforces 

existing statutes to prevent minority and low-income communities from 

being subjected to disproportionately high or adverse human health or 

environmental effects. 

The human health and environmental effects resulting from the preferred 

alternative are expected to be minimal, and are not expected to have 

disproportionate adverse effects to any minority or low-income family. 
IV. Environmental Consequences 97

Low-income families may depend on woodlots for firewood to heat their 

homes; however, the most valuable species used for firewood (including 

oak, hickory, beech, and locust) are not ALB-host species and would not 

be removed. Although some maple species may be less valued for 

firewood, they are commonly used for that purpose and are a preferred 

ALB-host. Nevertheless, if no action is taken, allowing ALB to spread 

could result in permanent loss of maples and all other ALB-hosts from the 

area. For full host removal, stumps from high-risk host trees in woodlots 

may be allowed to resprout, which would then allow more rapid regrowth. 

Wood treated with imidacloprid and used as firewood is not expected to 

cause adverse health effects. Therefore, the human health and 

environmental effects from the action alternatives are not expected to have 

disproportionate adverse effects to any minority or low-income family. 

i. Executive Order 13175 

EO 13175, “Consultation and Coordination with Indian Tribal 

Governments,” was issued to ensure that there would be “meaningful 

consultation and collaboration with tribal officials in the development of 

Federal policies that have tribal implications....” APHIS prepared and 

sent a letter to each of the tribes in the contiguous United States in June, 

2013, prior to publishing the NOI to prepare the EIS. The letter provided 

background on ALB in the United States, and provided the tribes an 

opportunity to dialogue with APHIS regarding potential impacts to tribal 

resources. APHIS also hosted a teleconference with all interested tribes in 

July, 2013, to discuss the Program and answer any questions from tribal 

representatives. APHIS will continue to actively engage with the tribes on 

eradication program activities, and address questions and concerns. This 

would occur through tribal consultation prior to the development of sitespecific EAs. 

j. Executive Order 13186 

EO 13186, “Responsibilities of Federal Agencies to Protect Migratory 

Birds,” directs Federal agencies taking actions with a measurable negative 

effect on migratory bird populations to develop and implement a 

memorandum of understanding (MOU) with FWS that promotes the 

conservation of migratory bird populations. On August 2, 2012, an MOU 

between APHIS and FWS was signed to facilitate the implementation of 

this EO. The MOU provides APHIS with guidance to avoid and 

minimize, to the extent practicable, detrimental migratory bird habitat 

alteration or unintentional take during management activities. 

k. Other Federal and State Regulations and Statutes 

Other Federal and State regulations may also apply to the proposed 

alternatives identified in this EIS. APHIS complies with all applicable 
98 IV. Environmental Consequences 

Federal regulations and Executive orders discussed in this EIS, such as 

the: 

 Clean Air Act, 

 Clean Water Act, 

 Coastal Zone Management Act, and 

 Federal Insecticide, Fungicide and Rodenticide Act. 

States may also have applicable regulations regarding various proposed 

activities related to the ALB Eradication Program. APHIS works 

cooperatively with State agencies in the implementation of any proposed 

ALB activities, and to identify applicable State regulations to ensure 

compliance. 

APHIS recognizes that the various alternatives proposed in this EIS may 

pose some risk to human health and the environment. This includes the no 

action alternative, as well as the proposed alternatives that were evaluated 

in this EIS. Many of the potential risks associated with implementation of 

an eradication program can be mitigated based on program requirements 

and, in the case of pesticide use, Federal regulations; this includes pesticide 

label language to reduce risk to human health and the environment. In 

addition, coordination with State agencies and other stakeholders on sitespecific eradication efforts ensures that other State and local requirements 

are considered and implemented, as appropriate. 

Hazards to the public from tree removal in urban and natural areas are 

mitigated through proper notification to ensure that the public would not 

be in areas when tree removal operations are taking place. All tree 

removals are coordinated with the respective landowners prior to any 

removal. Notification to landowners also occurs when chemical 

treatments may be used to kill stumps (herbicides) or treat high-risk host 

trees for ALB (insecticide). In addition to the notification process, there 

are other risk reduction requirements for pesticide use that are designed to 

protect workers, as well as the general public, and the environment. The 

mitigations are included on the pesticide label for a given formulation, and 

must be followed to ensure compliance with applicable Federal statutes. 

The greatest potential for human exposure to program pesticides is for 

workers who handle the concentrated form of each product and make 

applications. In the case of workers, there are several requirements for 

PPE that will reduce exposure and risk including the use of: 

 long sleeved shirt and long pants, 

 protective eyewear, 

 chemical resistant gloves, and 

 shoes and socks. 

2. Mitigation 

and Risk 

Reduction
IV. Environmental Consequences 99

These PPE requirements are based on the known toxicity profile for each 

of the proposed pesticides and the various formulations. Other labeling 

restrictions (e.g., applicable reentry intervals) are also designed to protect 

the public, including children as well as domestic pets. 

Risk reduction to the environment from the proposed alternatives may 

occur through program requirements, as well as pesticide labeling to 

restrict use. As an example of program measures to reduce risk, the 

current eradication activities in Ohio are implementing Ohio Department 

of Forestry BMPs that have been developed to protect soil and water 

quality during tree removal operations (Ohio Department of Natural 

Resources, n.d). These BMP’s identify stream management zones (SMZs) 

that are protected areas that have been established adjacent to bodies of 

water that increase with increasing slope (Ohio Department of Natural 

Resources, n.d). These types of buffers have been shown to provide 

protection to receiving waters from sedimentation and nutrients that may 

be of concern for a given watershed (Wenger, 1999). 

Excess sediments and nutrients are primary causes for listing waterways in 

the United States as impaired, under Section 303(d) of the CWA. The 

SMZs are areas where no, or very limited, cutting would be allowed. 

Infested ALB-host trees that could occur within a SMZ would still need to 

be selectively removed, but without the use of heavy equipment. This 

could also occur with some high-risk host trees; however, in those cases, 

the removals would only be conducted with landowner approval and 

consultation with other Federal and State agencies to ensure the risk to 

water quality is minimized. 

The risk to human health and the environment from the offsite transport of 

pesticides that are proposed for use in the ALB Eradication Program may 

also be reduced by pesticide labeling requirements designed to reduce the 

potential for contamination. Labeling restrictions and recommendations 

regarding application rate, timing of application, and application proximity 

to sensitive areas are all designed to reduce the risk of effects to human 

health and the environment. In addition, pesticide storage and disposal 

requirements are designed to reduce the risk of accidental spills and 

contamination of areas which could result in effects to human health and 

the environment. 

APHIS conducts various monitoring for several of its programs. 

Monitoring efforts are typically directed towards measuring potential 

exposure to program pesticide treatments. Monitoring may consist of grab 

samples collected from different matrices (e.g., soil, water, air, biota) 

based on potential concerns regarding exposure which may occur to the 

environment and human health. These monitoring efforts are conducted to 

support analysis and compliance with applicable Federal regulations, such 

as the NEPA, ESA, and Federal Insecticide Fungicide and Rodenticide 

3. Program 

Monitoring 
100 IV. Environmental Consequences 

Act. Monitoring efforts in previous and current ALB eradication 

programs have been directed towards analysis of imidacloprid in various 

matrices (APHIS, 2013a). The objectives of the ALB environmental 

monitoring program are to: 

 demonstrate the effectiveness of ALB-operational procedures in 

excluding or minimizing exposure of the public and the environment 

to Program-applied imidacloprid; 

 collect data which can be used to evaluate whether the assumptions 

used in the environmental assessments are valid estimates of potential 

exposure of the public and the environment to Program-applied 

imidacloprid; and 

 investigate any Program-related complaints or reports of adverse 

effects on public health, worker safety, environmental quality, or 

nontarget species. 

The APHIS environmental monitoring staff coordinates the collection, 

packaging, and shipment of samples to accredited laboratories for 

processing and analysis according to standard operating procedures 

(APHIS, 2013a). The results of the laboratory’s residue analyses are then 

correlated with environmental conditions data recorded at the time of 

treatment and sampling. The APHIS’ environmental monitoring staff 

further analyzes the results to determine whether there are any human or 

environmental risks related to the use of the pesticide. The monitoring 

staff reports the data and analyses to Program managers at the end of the 

program, or intermittently during the program, as required. 

ALB-related environmental monitoring for imidacloprid has occurred in 

Massachusetts and New York. To date, monitoring efforts for 

imidacloprid have focused on soil, water, plant, and bee-related samples to 

evaluate potential residues. Monitoring efforts will continue to be 

evaluated on a site-specific basis for any current and potential new 

infestations in the future. 

APHIS has been working on ALB eradication efforts since the beetle was 

first discovered in Brooklyn, New York, in 1996. Subsequently, resources 

have been committed to address infestations, as well as conduct research 

to understand the impacts of ALB introduction into the United States. 

Research to develop successful eradication methods have occurred, and 

will continue, so that APHIS implements the most effective methodologies 

in current and future infestations. Since the initial infestation, other 

infestations have been documented requiring additional APHIS resources. 

APHIS works with multiple stakeholders regarding the implementation of 

ALB-eradication activities. This includes cost-sharing to implement 

4. Irreversible 

and 

Irretrievable 

Commitment 

of Resources 
IV. Environmental Consequences 101

various aspects of the ALB Eradication Program. Federal share of the 

total costs to operate ALB eradication activities has varied from 45 to 

95 percent since the first Program was initiated (table 4–1). To date, 

Federal spending for the ALB Eradication Program has been over $500 

million. 

 

Table 4–1. ALB Eradication Program Historical Cost-Share Contributions 

(from 1997 to 2009 (thousands of dollars)).

Fiscal 

Year 

Federal 

Funding 

(Appropriated 

+ Emergency 

Funding) 

Cooperator 

Funding Total Federal Share 

1997 849 149 998 85%

1998 1,327 1,634 2,961 45%

1999 5,510 2,573 8,083 68%

2000 16,180 1,555 17,735 91%

2001 49,098 2,654 51,752 95%

2002 31,656 4,000 35,656 89%

2003 33,181 4,000 37,181 89%

2004 42,851 4,000 46,851 91%

2005 28,933 11,071 40,004 72%

2006 19,859 11,218 31,077 64%

2007 19,904 13,731 33,635 59%

2008 19,867 11,602 31,469 63%

2009 44,618 16,052 60,670 74%

2010 74,472 12,856 87,328 85% 

2011 32,456 9,843 42,299 77% 

2012 56,732 4,191 60,923 93% 

2013 39,731 2,240 42,151 94% 

Total $517,057 $111,599 $628,656 79%

 

 

 
102 IV. Environmental Consequences 

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Appendix A. Preparers A–1

Appendix A. Preparers 

U.S. Department of Agriculture 

Animal and Plant Health Inspection Service 

4700 River Road 

Riverdale, MD 20737 

 _____________________________________________________ 

James E. Warren 

Environmental Toxicologist 

B.S. Forest Management 

M.S. Entomology 

Ph.D. Environmental Toxicology 

Background: Environmental Protection Specialist in Environmental and 

Risk Analysis Services, with over 18 years of experience in environmental 

toxicology and risk assessment, as well as environmental fate modeling of 

pesticides while working for the Federal government and the agrochemical 

industry. 

EIS Responsibility: Project Lead for the Programmatic ALB Eradication 

Program EIS. Reviewed various chapters and contributed in writing 

various sections in the EIS, with an emphasis on the Environmental 

Consequences chapter. Coauthor for the chemical human health and 

ecological pesticide risk assessments. 

 Andrea Lemay 

Biological Scientist 

M.S. Plant Pathology 

Background: Biological scientist and plant pathologist in Environmental 

and Risk Analysis Services, with over 11 years of experience in risk 

analysis while working for the Federal government. 

EIS Responsibility: Analyst for the Programmatic ALB Eradication 

Program EIS. Reviewed various chapters and contributed in writing the 

Alternatives and Affected Environment chapters. 

 Robyn Rose 

National Policy Manager 

B.S. Agronomy 

M.S. Entomology 

Ph.D. Entomology 

Principal EIS 

Contributors 
A–2 Appendix A. Preparers 

Background: Domestic National Policy Manger in Plant Protection and 

Quarantine’s (PPQ) Plant Health Programs, with 17 years of experience as 

an environmental risk assessor and program manager. 

EIS Responsibility: Providing technical expertise regarding Programmatic 

ALB Eradication Program EIS and the program, and serving as the 

primary reviewer for PPQ. 

 Vivian A. Miller 

Environmental Protection Specialist 

B.S. Biology 

M.Ed. Secondary Education, Biology 

Background: Environmental protection specialist with over 4 years of 

experience preparing and reviewing NEPA compliance documents for 

APHIS. 

EIS Responsibility: Analyst for the Programmatic ALB Eradication 

Program EIS. Reviewed various chapters and contributed in writing the 

Affected Environment and Environmental Consequences chapters. 

 Tracy A. Willard 

Environmental Protection Specialist 

B.S. Biology 

M.S. Entomology 

Ph.D. Entomology 

Background: Fourteen years of service with APHIS preparing 

environmental documents. Experience in environmental compliance, 

especially as associated with the Endangered Species Act. 

EIS Responsibility: Analyst for the Programmatic ALB Eradication 

Program EIS. Prepared chapter 1 and reviewed various chapters. 

 Fan Wang-Cahill 

Environmental Health Specialist 

B.S. Biology 

M.S. Hydrobiology 

Ph.D. Botany 

Background: Eighteen years professional experience in human health risk 

assessment for environmental contaminants at Superfund, Resource 

Conservation and Recovery Act, and State-regulated contaminated 

facilities. Expertise in preparing human health risk assessments for 

APHIS. 
Appendix A. Preparers A–3

EIS Responsibility: Analyst for the Programmatic ALB Eradication 

Program EIS. Contributed in writing the human health effects of the 

environmental consequences chapter. 

 

Allan N. Auclair 

Systems Ecologist 

B.S. General Science 

Ph.D. Plant and Systems Ecology 

Background: Risk scientist with expertise on pest risk assessment, pest 

outbreak dynamics, and the role of climate drives and climate change on 

pest incidence. Eight years of service in APHIS. Experience in risk 

analysis especially as it relates to risk assessment of exotic pests entering 

on commodities under import permit application; climate as a drive on the 

outbreak of high-threat pests; and compliance to climate change and 

invasive pest legislation including Executive Orders 13122 and 13514 and 

Departmental Regulation 1070–001. 

EIS Responsibility: EIS analyst – evaluated the effects of removing ALBinfested trees on greenhouse gas (GHG) emissions, and the effects of 

climate change impacts on/by tree removal actions. Identified 

management alternatives and “preferred option” to minimize impacts. 

 

Eleni C. Tsigas 

Agricultural Economist 

B.S. Mathematical Economics 

Ph.D. Agricultural Trade and Policy 

Background: Applied economist analyzing the economic impact(s) from a 

proposed change in an agricultural policy/rule (domestic or foreign) and 

presenting the decision makers with the outcome of the potential economic 

benefits and costs for all the participants. Sixteen years of service with 

APHIS’ Policy Analysis and Development Programs Division. 

Worked as an analyst with the International Markets Division in Economic 

Research Service of USDA, and in the International Food and Policy 

Research Institute (IFPRI) for 7years. 

EIS Responsibility: Economic analyst 

 
A–4 Appendix A. Preparers 

 (This page is intentionally left blank.) 
Appendix B. List of ALB NEPA Links B–1

Appendix B. List of ALB NEPA Links Inventory of 

USDA Analyses to Date with Links to the 

Resources 

 Asian Longhorned Beetle Cooperative Eradication Program in Clermont County, Ohio, Revised 

Environmental Assessment, May 2013 (PDF; 2.65 Mb) 

 FONSI signed May 1, 2013 (PDF: 71 Kb) 

 

Asian Longhorned Beetle Cooperative Eradication Program in Clermont County, Ohio, 

Environmental Assessment, May 2012 (PDF; 1.27 Mb)

Asian Longhorned Beetle Eradication Efforts in Clermont and Brown Counties, Ohio, 

Environmental Assessment, September 2011 (PDF; 3.03 Mb) 

 FONSI signed September 6, 2011 (PDF; 103 Kb) 

Asian Longhorned Beetle Cooperative Eradication Program in Essex, Norfolk and Suffolk 

Counties, Massachusetts, Environmental Assessment, May 2011 (PDF; 83 Kb) 

 FONSI signed May 20, 2011 (PDF; 4.28 Mb) 

Chemical Treatment Study in New York City, New York, and Central New Jersey for the Asian 

Longhorned Beetle Eradication Program Environmental Assessment, September 2010 (PDF; 

5.10 Mb) 

 FONSI signed September 14, 2010 (PDF; 93 Kb) 

Nursery Treatment Efficacy Study within Worcester County, Massachusetts, to Support the 

Asian Longhorned Beetle Cooperative Eradication Program, Environmental Assessment, April 

2010 (PDF; 192 Kb) 

 FONSI signed May 6, 2010 (PDF; 1.07 Mb) 

New Chemical Treatment Study within the Worcester, Massachusetts, Quarantine Zone for the 

Asian Longhorned Beetle Eradication Program, Environmental Assessment, September 2009

(PDF; 903 Kb) 

 FONSI signed October 30, 2009 (PDF; 162 Kb) 

Asian Longhorned Beetle Cooperative Eradication Program in Worcester and Middlesex 

Counties, Massachusetts Environmental Assessment September 2008 (PDF; 1.28 Mb) [FONSI 

signed 11/21/08] 

 Amended FONSI signed September 11, 2009 (PDF; 140 Kb) 
B–2 Appendix B. List of ALB NEPA Links 

 Addendum to the FONSI signed March 29, 2010 (PDF; 469 Kb) 

Asian Longhorned Beetle Cooperative Eradication Program in the New York Metropolitan Area, 

Environmental Assessment May 2007 (PDF; 43 Kb) 

Asian Longhorned Beetle Cooperative Eradication Program - Hudson County, New Jersey, 

Environmental Assessment, March 2003 (PDF; 36 Kb) 

Asian Longhorned Beetle Program, Environmental Assessment, February 2000 (PDF; 892 Kb) 

Asian Longhorned Beetle Field Trial, Environmental Assessment, January 2000 (PDF; 698 Kb) 

Asian Longhorned Beetle Control Program - Illinois, Environmental Assessment, August 1998

(PDF; 23 Kb) 

Asian Longhorned Beetle Control Program, Environmental Assessment, December 1996 (PDF; 

23 Kb) 
Appendix C. ALB Annotated Host List C–1 

Appendix C. ALB: Annotated Host List

Updated by Baode Wang, January 2015 

USDA-APHIS-PPQ, Center for Plant Health Science and Technology, Otis Laboratory 

Last accessed May 15, 2015 at 

http://www.aphis.usda.gov/plant_health/plant_pest_info/asian_lhb/downloads/hostlist.pdf

Genus1 Common Name Host Abundance and Other Notes2 Treated, 

Surveyed3

Preferred host in US4

Acer Maple, boxelder Very common trees. Many US records, all species: Norway, red, silver, sugar, 

sycamore maple and boxelder especially favored; Amur maple less favored; 

Japanese maple seldom attacked. 

yes

Aesculus Horsechestnut, buckeye Fairly common trees. Several US records, some heavily infested. yes

Betula Birch Fairly common trees. Several US records: gray, European white and river 

birches. Some gray birches with many exits. Birches are apparently less 

preferred than maple. No exit holes found in laboratory studies with black and 

yellow birches yet although some larva developments inside trees of these two 

species have been observed. 

yes

Salix Willow Fairly common trees. Several US records: weeping, pussy and white 

willows highly favored; black willow (oviposition only) less favored. yes

Ulmus Elm Very common trees. Many US records: American, Siberian and Chinese elms. 

Elms are apparently less preferred than maple. yes

Occasional to rare host in US4

Albizia Mimosa, silk tree, 

A. julibrissin

Occasional ornamental. Exit holes: 2 records from field in NY with 

additional emergence in laboratory. No Chinese record. yes

Cercidiphyllum Katsura tree, 

C. japonicum

Occasional ornamental. Four records from Worcester, MA, including 2 trees 

with exit holes. yes

Fraxinus Ash (especially 

green ash, F. 

pennsylvanica) 

Very common tree, but injury infrequent relative to host abundance. Several 

US records, all from IL, most of these unverified (but at least two exit holes 

confirmed). Chinese ash, F. chinensis and white ash, F. americana were 

confirmed to be host in China

yes

Platanus London plane tree, 

P. acerifolia

Very common urban trees. 12 US records (including 4 with exit holes, NY); 

no record for P. occidentalis, American sycamore. Host in Chinese literature. 

Exit holes observed in China. 

yes

Populus Poplar Very common trees. Diverse and variable group, hybrids occur. Suitability 

apparently varies; some species and hybrids are prime hosts in China, others 

are rare host. Nine US records (NY, NJ, MA). Complete life cycle on eastern 

cottonwood, P. deltoides and quaking aspen, P. tremuloides. Oviposition on 

balsam popular, P. balsamifera, Balm-of- Gilead (a hybrid cultivar), 

unidentified Populus sp. Generally, Populus section Aigeiros (black poplars) 

are more preferred than other sections.

yes

Sorbus European mountain-ash, 

S. aucuparia

Occasional ornamental. Exit hole: 1 record from field in IL with additional 

emergence in laboratory. No Chinese record. Note: this is not a true ash; 

Sorbus is a member of the rose family. 

yes
C–2 Appendix C. ALB: Annotated Host List 

Genus1

Common Name Host Abundance and Other Notes2 Treated, 

Surveyed3

Questionable US records4

Celtis Hackberry, 

C. occidentalis

Fairly common tree. Oviposition: 1 unverified record from IL, with 

small/medium-sized larva identified as ALB. No Chinese record. No egg sites 

were found in laboratory studies with caged trees and beetles and no active egg 

sites or exit holes were found in ALB host studies in a “common garden” setting 

and surveys in China. Feeding by adults was observed. 

no 

Hibiscus Rose-of-Sharon, 

H. syriacus

Common ornamental shrub. Exit: 1 unverified report, NY; Oviposition: several 

records, NY, but no larval development, possibly incidental to heavy damage on 

nearby hosts. No Chinese record. Adult feeding, oviposition, egg sites and 

active egg sites were observed in caged studies in “common garden” settings in 

China

no 

Malus Apple, crab apple Common ornamental. Oviposition: 1 questionable record, IL. Host in Chinese 

literature. Oviposition observed in China. No exit holes found yet. no 

Morus Mulberry Very common tree. Oviposition: 1 record, NY. No Chinese record. Unlikely 

to be ALB host. no 

Prunus Cherry, plum Very common ornamental. Oviposition: 2 records, NY & IL, but no survival. 

Host in Chinese literature. No exit holes have been found in our study in 

“common garden” setting. 

no 

Pyrus Pear Common ornamental. Exit: 1 questionable record, IL. Host in Chinese literature. 

Few exit holes were observed on Pyrus bretschneideri trees in China. no 

Quercus Oak, 

(pin oak, Q. palustris) 

Very common tree. Oviposition: 1 record, NY (incidental to heavy damage 

on nearby hosts). No Chinese record. no 

Robinia Black locust, 

R. pseudoacacia

Common tree. Exit: 2 doubtful records, IL. Host in Chinese literature. Quite a 

few egg sites were observed in China, no exit holes. no 

Tilia Linden (littleleaf linden, T. 

cordata) 

Common tree. Oviposition: 2 records (IL & NY) but no survival. 

Oviposition but no survival in Canada. Host in Chinese literature. no 

No US record4

Alnus Alder Locally common tree or shrub. No US record. Host in Chinese literature. Exit 

hole observed in gray alder, A. incana, in caged study in China. no 

Elaeagnus Russian live 

(Oleaster), E

angustifolia

Widely-distributed ornamental shrub and escaped weed; quite variable, easily 

confused with other Elaeagnus species. No US record. Host in Chinese 

literature; Heavy feeding damage and few exit holes observed in China. 

no 

Koelreuteria Goldenraintree, K. 

paniculata

Occasional ornamental. No US record. Heavy feeding, oviposition sites and 2 

exit holes observed in field studies in China. Other exit holes were also found 

on trees along roadside. 

Yes 

Melia Chinaberry, 

M. azedarach

Uncommon shrub. No US record; reported not to be a host in Chinese 

literature but damage observed. Host of the citrus longhorned beetle, 

Anoplophora chinensis. 

no 

Non-host4

Ailanthus 

 Tree of heaven, 

A. altissima

Common tree. No US record; reported not to be a host in Chinese 

literature. no 
Appendix C. ALB Annotated Host List C–3 

1. Host genera listed alphabetically within categories. 

2. Host abundance based on (a) records and observations of infested areas in NY, IL, NJ and 

MA, (b) Nowak (1994) and (c) descriptions of range and abundance in several field guides. 

3. Included in surveys and chemical treatments by USDA Cooperative ALB Eradication 

Program in IL, NY, NJ and MA. 

4. Host status based on US records of infestation, field studies with North American trees 

planted in China and Chinese literature. Host range tests in laboratory and greenhouse 

settings not considered except as noted. See Hu et al. (2009) for a review of hosts with 

particular emphasis on the status of poplars in China. 

Additional notes: 

1. Celtis occidentalis is most likely not a host of ALB, field studies, surveys and observations in 

China have found no evidence of Celtis as ALB host. However, its status would change if 

surveys reveal any infestation. 

2. Styphnolobium japonicum syn. Sophora japonica (the pagoda tree), ALB completed 

development in 2-3 years on this species in caged study. 

3. Rosa listed as host in some literature. The beetle may be the citrus longhorned beetle (CLB), 

Anoplophora chinensis. Some species of trees in Rosa are good host of this beetle. 

References 

Hu, J., S. Angeli, S. Schuetz, Y. Luo and A. E. Hajek. 2009. Ecology and management of exotic 

and endemic Asian longhorned beetle Anoplophora glabripennis. Agric. For. Entomol. 11: 359-

375. 

Nowak, D. J., 1994, “Urban Forest Structure: The State of Chicago’s Urban Forest,” pp. 3-18 In: 

E. G. McPherson et al., Chicago’s Urban Forest Ecosystem: Results of the Chicago Urban Forest 

Climate Project. Gen. Tech. Rep. NE-186, USDA Forest Service, NE Forest Experiment Sta., 

Radnor, PA. 

 
C–4 Appendix C. ALB: Annotated Host List 

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Appendix D. Climate Change Impacts Related to ALB D–1 

Appendix D. Assessment of Greenhouse Gas Fluxes 

and Climate Change Impacts Related 

to ALB 

 
D–2 Appendix D. Climate Change Impacts Related to ALB

Table of Contents

EXECUTIVE SUMMARY 

1.0 INTRODUCTION 

2.0 GHG EMISSIONS AND PROJECT IMPACT ON CLIMATE CHANGE 

3.0 ALTERNATIVES 

4.0 MITIGATION MEASURES 

5.0 REFERENCES ATTACHMENTS 
Appendix D. Climate Change Impacts Related to ALB D–3 

Executive Summary 

 

The five key findings in this assessment of Asian longhorned beetle (ALB) eradication impacts 

are: 

1. Total carbon dioxide (CO2) emission from trees and soil at the Clermont County Ohio ALB 

Eradication Project site is estimated to be 7,872 metric tons (mt); total at all five ALB 

eradication sites is estimated to be 20,187 mt. These levels are below the CEQ (Council on 

Environmental Quality) reference level of 25,000 mt for all greenhouse gases (GHG); other 

GHGs (methane, nitrous oxide, hydro-fluorocarbons, perfluorocarbons, and sulfur hexafluoride) 

were not assessed in this study. Alternative scenarios of ‘no action’ indicate 2,254 million mt of 

CO2 would be released from host trees in urban areas and adjacent forests under the current tree 

removal protocols across 48 States of the contiguous US. In the extreme scenario where all host 

tree species are infested by ALB across the 48 States, 13,906 million mt would be released in 

the absence of any eradication. 

2. Cumulative CO2 emission is zero by 76 years post-eradication where trees removed are 

artificially re-planted and/or natural forest regeneration is allowed. Most cities in the US this 

decade show diminished tree cover due to increased infrastructure (i.e., road, residential, and 

commercial building). Compared to American Forests standards of 40% urban tree cover, the 

US-wide average is 23%, indicating considerable leeway to increase CO2 sequestration and C 

storage through expansion of urban tree cover. Alternative scenarios indicate that under 

methods that enhance tree growth by 10%, provide added protection, and expand tree cover 

from 25% to 40%, complete CO2 recovery could be achieved in 43 years and total C storage 

could double over ‘business-as-usual’ approaches. 

3. Tree removal in urban areas and adjacent forest impacts negatively on building energy 

consumption, local air pollution levels, and flooding control. Promoting tree cover to non- host, 

rapidly growing tree species increases adaptation to climate change. Examples are given of US 

cites with ‘greening’ programs aimed at increased tree cover to achieve benefits of adaptation. 

4. Both host trees and ALB are strongly impacted by present and continuing climate change. 

In a study of Northeast US forests, 36 out of 80 species assessed show potential for their 

ecological optima to shift at least 100 km to the north, including seven that could move >250 

km. Sugar maple (Acer saccharum, -14.5%) and black cherry (Prunus serotina, -10.0%) would 

decline sharply, while oak and southern pines would expand northward – hence an expected 

shift away from prime ALB host trees in the near future. ALB under present climate has the 

potential to occupy all 48 contiguous States. Temperature increases would enhance ALB 

survival and growth over large areas. Data on a diverse array of 612 crop pests and pathogens 

worldwide demonstrate an average pole-ward shift of 2.7 +/-0.8 km / year since 1960, 

supporting the hypothesis that global warming-driven pest movement is already underway. 

5. Long-term shifts in average temperature and total precipitation at Cincinnati, Ohio (1895- 

2013) are modest, highly variable, and cyclical; temperature increased 0.42C/100 years 

compared to a global temperature change 0.73C/100 years. A simple ALB ‘Index of Climate 

Favorability’, 1950 to 2013 at Cincinnati is dominated by a cyclic pattern. Linear regression 

shows a decrease rate of 1.6 %/100 years. The present outbreak in Clermont County Ohio is 

preceded by a 6-year interval of favorable climate for ALB; since first detection in1996. 
D–4 Appendix D. Climate Change Impacts Related to ALB

Years 1996, 2000, 2001, 2003, 2004, 2009 stand out as particularly favorable for ALB growth; this 

contrasts to any earlier 18-year period which had only two to three such events. Index values 

suggest an overall static trend but one punctuated by especially favorable or adverse yearly 

conditions. Outbreak sites imply that ALB thrives within a ‘humidity corridor’ and that commonly 

applied degree-day models require revision to fully incorporate moisture parameters. 

 

1.0 Introduction

USDA APHIS evaluated the potential for greenhouse gas (GHG) emissions related to taking no 

action against the Asian longhorned beetle (ALB) and the preferred alternative which includes 

removal of infested host trees and a combination of removal and imidacloprid treatments of high 

risk host trees. 

 

Estimating these types of emissions can be difficult to do on a national scale especially in this case 

where ALB has not been detected in areas where it could become established but could be in the 

future due to the presence of host trees and favorable environmental conditions. Estimates 

regarding the release of GHG in this assessment were quantified based on the most recent ALB 

outbreak in Ohio with information also provided on the other ALB outbreaks that have occurred in 

the United States. The emission of GHGs associated with tree removal, and GHG sequestration 

where tree replanting and regeneration are allowed (albeit into different tree species) was evaluated 

against the tree and soil carbon balances typical of similar but non ALB- infested areas. 

 

2.0 GHG Emissions and Project Impact on Climate Change 

2.1 GHG Emissions and Sequestration

2.1.1 Direct and indirect GHG emissions

 

Direct Emissions. The most conspicuous impact of the removal of large numbers of ALB- infested 

trees is the loss of CO2 sequestration and release of carbon storage in trees and soil. As of mid2013, about 9,400 trees had been removed in the Ohio ALB eradication program (Table 1). 

 

Levels of tree removal at five ALB eradication sites were used as input to a CO2 Emissions and 

Sequestration Model. This model developed by us applied the number of trees removed in tree 

species biomass equations of the USDA Forest Service, assumptions of tree size distribution among 

urban and forest tree populations, tree carbon and soil carbon content to estimate total CO2 emitted 

to the atmosphere through decomposition of wood, and that sequestered from the atmosphere in 

regrowth over tree life expectancy of 100 years (Attachment 1). 

Release of CO2 to the atmosphere through decomposition follows removal, cutting, and chipping. 

Rapid release initially occurs from fine, nutrient-rich branches; this is followed by breakdown of 

residual heartwood and larger branches, resulting in a transient spike of release 8-12 years later. 

Decomposition in most temperate hardwoods is more or less complete by Year 22 (Fig. 1). 

 
Appendix D. Climate Change Impacts Related to ALB D–5 

Assuming artificial replanting and natural regeneration follows immediately after tree removal, 

sequestration of CO2 in new growth will match that released annually in decomposition by Year 16 

and will completely re-capture all the CO2 released by Year 76. Hence, the net positive flux (to 

atmosphere) in early decades reverts to a net negative flux (from atmosphere) in later decades. This 

changing dynamic is an important element in the overall, complete assessment of CO2 impacts on 

GHGs (Fig. 1). Assertive management of tree health and expanded tree cover (as discussed below 

under Alternative 2) can actually benefit in shortening the time to full carbon recovery, and may 

substantially exceed the initial losses through better tree growth and through the selection of 

replacement species both resistant to ALB and better suited to warming conditions and associated 

weather variability. 

The Ohio site to date is estimated to have released a total of 1762 mt of CO2, about a quarter of that 

from soil disturbance; eventually that level could rise to 7872 mt. The total eventual release of CO2 

from all five ALB eradication sites in the northeast and north central States is estimated to be 

20,200 mt through the near future, based on removal estimates of about 114,000 trees (Table 1). 

Figure 1. Estimates for Ohio ALB eradication program of the emission of CO2 from trees removed 

(blue) and sequestration of CO2 in new tree growth (green). Net CO2 flux (red) is emissions minus 

sequestration; dashed red line is zero net flux. Full recovery of all initial CO2 emission is captured 

in new tree growth (dashed blue line, Year 76). 

 
D–6 Appendix D. Climate Change Impacts Related to ALB

Table 1. Summary of tree removal levels through July 2013 and estimated final number at five 

locations in Northeastern United States, part of USDA APHIS Asian Longhorn Beetle Eradication 

Program. Estimated percent urban/forest, CO2 released following tree removal and chipping, years 

required for total CO2 release to atmosphere, and years required for total CO2 recovery through 

sequestration in new tree growth. 

 Indirect Emissions. Other GHGs tied to the carbon cycle, notably NO2 and CH4 can have 

important atmospheric warming effects but were not quantified in this study. Methane 

production following tree removal would occur where anaerobic conditions exist, such as in very 

wet or submerged soils. Nitrogen compounds that eventually may end up in the atmosphere as 

GHG are released as tree wood decomposes. The production of nitrates (- NO3), studied in-depth 

by researchers more than other forms of organic nitrogen, is closely tied to soil processes and 

stream water discharge. Christopher et al (2004) observed that – NO3 discharge from forested 

watersheds in New York was largely from surface soils following snowmelt and major storm 

events (i.e., the ‘flushing effect’). Even though the geology and hydrology of the two adjacent 

deciduous watersheds they studied were similar, the levels of –

NO3 in discharge showed 

significant differences -- for reasons likely attributed to dissimilar nitrogen-fixing and 

decomposition properties. The lack of ALB Program data on soils and soil disturbances during 

cutting operations makes meaningful estimates of methane and nitrate production/release to the 

atmosphere difficult. In the eastern US, atmospheric nitrogen deposition is high, complicating an 

overall assessment of nitrogen release due to tree removal. 

The USDA OCE (2013) recently released a draft report on Science-Based Methods for EntityScale Quantification of Greenhouse Gas Sources and Sinks from Agriculture and Forestry 

Practices. Nitrogen and methane estimates form part of the agricultural methods documented, 

but do not appear in any of the forest or tree-based management models (e.g., forestry, 

afforestation, agro-forestry, wood-carb biomass). 
Appendix D. Climate Change Impacts Related to ALB D–7 

2.1.2 GHG emissions in carbon equivalents

The only GHG quantified in detail is CO2. Figure 1 and Table 1 are representative estimates of 

CO2 release and sequestration for the Ohio and four other ALB eradication projects to date. 

Throughout, we used the commonly reported conversion factor of 3.67 grams CO2 per gram of 

carbon. 

2.1.3 Cumulative impacts project would contribute to/have on global 

climate change

The estimated total CO2 release from five ALB eradication sites in the northeast US is 20 

thousand mt (Table 1). This is approximately 0.024 % of total global CO2 annual emissions 

based on the Year 2010 estimate of 9.1 Gt C / year (33.5 Gt CO2) global emissions from 

industrial sources and an estimated total of 10.0 Gt C / year, including land-use change and 

deforestation (http:// co2now.org/Current-CO2/CO2-Now/global-carbon-emissions.html 

accessed 10/30/2013). 

Unlike industrial emissions from fossil fuel sources, forests have the advantage of carbon 

recovery. Through sequestration in new growth, natural and artificial regeneration would recapture the equivalent of the CO2 released within 76 years. Under improved management of 

urban tree health and tree cover, the level of CO2 sequestration in new tree growth could exceed 

initial losses due to tree removal and recapture fully all CO2 within as little as 43 years (see 

Alternative 2 below). 

2.2 Impacts of Project on/by Climate Change

2.2.1 Direct and indirect impact of climate change on the ALB 

eradication project

Long-term change in climate will vividly impact on both the (a) host trees (in urban settings as 

well as in natural forests) and on (b) ALB survival and population vigor (and that of most other 

invasive forest pests). 

(a)Direct impact of climate change on host trees and on vegetation. 

Anticipated changes in the New York City, Chicago, and Toronto urban areas (Nowak 

et al 2007, Nowak et al 2010, Toronto 2007) include: 

• Warmer winter temperatures and longer growing seasons; 

• Changes in the seasonality of precipitation and extreme events like droughts 

and heavy rainfalls; 

• Expanded ranges of insects and increased over-winter survival rates; 

• Increased frequency and severity of storm events. 

 

Model projections for the northeastern US show a range of impacts on forest 

productivity due to anticipated climate change. The overall effect on net ecosystem 
D–8 Appendix D. Climate Change Impacts Related to ALB

productivity is positive (+75 gC/m2/year but with notably large variations: from -150 to 

350 gC/m2

/year for the region, and -85 to +275 gC/m2/year at the level of four specific 

sites) (Aber et al 1995). Tree growth predictions vary because of the intermixing of 

positive and negative climate effects. Increases in carbon dioxide and temperature may 

have a positive effect by increasing the rate of tree growth, but only up to a point. 

Increased temperatures will also increase evapotranspiration, soil drying, and the 

frequency of short-term droughts, which would limit water availability for tree growth 

(Wilmont 2011). Overall, drops in tree productivity and associated changes in microclimate imply stress and vulnerability of trees to pathogen and insect attack. 

The expected shift is also toward more damaging events: increases and/or altered 

frequencies are expected for fire, drought, insect and pathogen outbreaks, prevalence of 

introduced species, hurricanes, windstorms and ice-storms (Dale and Beyeler 2001). 

Important to the issue in ALB management are large differences among major forest 

species in how and to the extent they will be affected. For example, for northeast US 

forests, Iverson and Prasad (1998) show that roughly 30 species could expand their 

range and/or weighted importance at least 10%, while an additional 30 species could 

decrease by at least 10%. Depending on the global change scenario used, 4–9 species 

would potentially move out of the United States to the north. Nearly half of the species 

assessed (36 out of 80) showed the potential for the ecological optima to shift at least 

100 km to the north, including seven that could move >250 km. They chose Vinton 

County in southern Ohio to model climate projections -- among 15 species projected to 

decline, sugar maple (Acer saccharum, -14.5%) and black cherry (Prunus serotina, -

10.0%) would decline sharply, while sourwood (Oxydendrum arboretum, +7.5%) and 

white oak (Quercus alba, +4.0%) would increase in importance. An additional 14 

species are projected to change very little under the changed climate. Hence, a major 

ALB host is projected to be one of the tree species in the northeast US most affected 

under expected near-future climate shifts. 

Overpeck et al (1991) predict the above general patterns for northern pines, as well as 

large increases in oak abundance in the northern Great Lakes and New England -- 

namely for black, northern red, and white oak (Quercus velutina, Q. rubra, and Q. 

alba,respectively). They also predicted a severe northern shift for white birch (Betula 

papyrifera), and a large northward expansion for southern pines, as exemplified by 

loblolly pine (Pinus taeda). The significance here is that there is an expected shift away 

from prime ALB host trees in the future. 

A recent case study of impacts on Vermont forests is particularly instructive of the 

degree of on-going and already visible changes within the current geographic reach of 

ALB. Wilmont (2011) notes that climate changes are already clearly evident in in the 

State. Temperatures have increased in the larger Northeast region by 1.8F (1.0C) since 

1970, with winter temperatures rising faster than summer temperatures. Precipitation has 

increased by 15-20% over the past 50 years with 67% of this now falling in heavy 

precipitation events. These and other medium-term trends in climate are anticipated to 

affect Vermont’s forests, including: 

 
Appendix D. Climate Change Impacts Related to ALB D–9 

• More frequent hot (over 90oF), humid days; 

• Longer growing seasons; 

• Worsening of air quality in areas where air quality problems already exist; 

• Increased heavy downpours; 

• More frequent winter thaws and earlier springs; 

• Less winter precipitation falling as snow and more as rain; 

• Earlier spring snowmelt resulting in earlier peak river flows; and, 

• More frequent short-term droughts in late summer and fall. 

Species distribution has shifted at high elevations in Vermont in the past 40 years. 

Northern hardwood trees are now able to survive at increasing elevations due to 

moderating temperatures, outcompeting spruce and fir trees which themselves are 

increasingly vulnerable to warming. Only slightly less vulnerable are northern hardwood 

forests whose dominant species are sugar maple, yellow birch and American beech. 

These forests are expected to be nearly eliminated in Vermont, replaced by species that 

prefer the warmer drier conditions, such as oak and pine species (Karl et al, 2009). Trees 

stressed from low water availability tend to reduce their defense mechanisms and are 

more susceptible to insect or disease invasion. Currently, three non-native pests are 

expanding across Vermont forests, with the potential to severely impact hemlock, ash, 

and maple species (Wilmont, 2011). 

Indirect impact of climate change on host trees and on vegetation. Elsewhere, such as in 

Britain, surveys of forest tree condition illustrate correlations between drought years and 

tree crown defoliation as drought, combined with high temperature induces stress in 

trees and thereby predisposes them to attacks by pests and pathogens. Such stresses are 

likely to be most intensely felt by street trees because of the ‘heat island effect’ and 

also because urban trees are often planted in suboptimal conditions with little area for 

root expansion and frequent root disturbance from utilities operations. This stress is 

often found to increase susceptibility to insect pests although evidence is more limited 

or at least variable, when it comes to pathogens. Changes in the plants’ environment 

have the capacity to alter their palatability to pests, as they can change the way in which 

plants allocate available resources to growth or defense. For example, increased CO2 

levels can make plants more palatable to some pest species through increases in soluble 

carbohydrates within phloem tissue (Tubby and Webber 2010). 

(b) Direct impact of climate change on Asian longhorned beetle. We anticipate that ALB 

and other insects will be strongly impacted by projected levels of global (and local) 

warming. Writ-large, predicted climate changes are likely to increase developmental 

rates and reduce winter mortality for many insects, leading to multi-voltinism in some 

species. The increased number of generations per year will also enable pests to evolve 

and adapt much more effectively to climatic change than their tree hosts, through 

enhanced dispersal as well as phenotypic and genotypic plasticity. 

Prior to Bebber et al (2013), the extent to which crop pests and pathogens alter their 

latitudinal ranges in response to global warming was largely unknown. Using CABI 

(Centre for Agriculture and Biosciences International) data on a diverse array of 612 

crop pests and pathogens worldwide, they demonstrate an average pole-ward shift 
D–10 Appendix D. Climate Change Impacts Related to ALB

of 2.7 +/- 0.8 km / year since 1960, but with significant variation in trends among 

taxonomic groups. Insect pests are strongly influenced: warming generally stimulates 

insect herbivory at higher latitudes, primarily through increased winter survival; rainfall 

has also an obvious effect, with defoliating insects responding negatively to drought and 

borers positively. The observed pole-ward shift in many taxa support the hypothesis that 

global warming-driven pest movement is already underway. 

Because of their generally short life-cycles, great reproductive potential, sensitivity to 

changes in temperature and, in many cases, great capacity for dispersal, even moderate 

changes in climate have already had significant rapid impacts on the distribution and 

abundance of many pests and pathogens. Climate change will also affect pest vectors, 

parasites and natural enemies. Native pests, some not currently perceived as problematic, 

may become more damaging. In comparison to insects, trees are very long-lived and will 

only be able to adapt much more slowly to changes in their local environment. 

Potentially, this makes them more vulnerable to the rapidly moving and changing 

organisms around them. 

There are few studies on the effect climate change will have on ALB in North America. 

On the other hand, there are several comprehensive laboratory studies of temperature and 

humidity impacts on the species’ reproduction by Keena and colleagues (Keena, 2006; 

2009; Keena and Moore, 2010; Sanchez and Keena, 2013); these provide much insight 

into the insect’s likely response to warming conditions. 

At an early point in her ALB research, Keena (2006) noted there was a critical need for 

information on the basic biology of the ALB. Accurate data and models are seen as 

essential underpinnings for scientifically predicting the insect’s development – with the 

management goals of optimizing exclusion and eradication treatments and predicting 

spread rates under different environmental conditions. Detailed studies followed on 

individuals from Bayside, NY and Ravenswood, IL to assess temperature effects on 

developmental rates and survival over a wide temperature range -- from 10°C (larvae 

only), 15°C, 20°C, 25°C, 30°C (egg and larvae only) to 35°C (larvae only). 

 

Keena and colleagues (Keena, 2006; 2009; Keena and Moore, 2010; Sanchez and Keena, 

2013) found: 

 

 a nearly linear relationship exists between developmental rate of ALB eggs and 

temperatures between 15°C and 30°C. Using this relationship, they predict that eggs 

would not hatch at temperatures of 10°C or less. Based on the lower percentage hatch 

of viable eggs at 30°C compared to 25°C, the upper temperature at which egg 

development ceases and eggs die is at or above 35°C. 

 

 defining lower and upper temperature limits are essential to accurate modeling. The 

minimum developmental threshold for instars 1 to 8 is close to 10°C. There is some 

development at 10°C, at least for early instars; about 20 percent of the larvae will 

molt to the second instar after about 5 months. The upper threshold at which 

development stops and death occurs, is probably between 35 and 40°C. Larval 

survival was higher at 25°C than at lower temperatures and was zero for 35°C by the 
Appendix D. Climate Change Impacts Related to ALB D–11 

beginning of the fifth instar. Larvae held at 15°C, 30°C, and 35°C had narrower head 

capsules and weighed less than those held at 20°C or 25°C from the third instar. 

 temperature response depended on ALB provenance. Bayside NY females laid fewer 

eggs at both 20°C and 25°C than Chicago IL females. The percentage of eggs that 

were viable did not vary between temperatures or strains. The Chicago IL larvae 

gained weight faster than those from Bayside NY in later instars. Temperature and 

its influence on larval weight had profound impacts on whether a larva proceeded to 

pupation. 

 there is no significant difference in female longevity between populations of the two 

provenances or temperatures. Males tended to lived longer than females at all 

temperatures and lived longer at 20°C than 15°C or 25°C. The time the females 

began laying eggs and the order of average number of eggs laid was significant: 

25°C > 20°C > 15°C. 

 the estimated lower threshold temperature for development of instars 1-5 and the 

pupal stage was near 10C and was near 12C for the higher instars. Developmental 

rate was less temperature sensitive for instars 5-9 compared with instars 1-4. 

Development for all but the first instar was inhibited at constant temperatures at 

about 30C, and all instars failed to develop at 40C. 

 differences in humidity shorten adult longevity, especially at low and high 

temperatures. Hence, caution should be used in applying predictions based only on 

temperature relationships developed in the laboratory to field conditions. 

Moreover, the beetle’s ability to seek out locations with optimum temperatures 

(e.g., sunny perches when it is too cool and shady locations when it is hot) may 

lessen the adverse effects in both the summer and fall; at least 2o

C should be added 

to air temperatures to adjust for the mediation of temperature by the wood. 

In summary, temperature has significant impact on all life history parameters (esp. female 

fecundity, longevity) assessed and comparable with that Zhou et al (1984) found for A. 

glabripennis (form nobilis) adults in China. Straightforward mathematical relationships greatly 

facilitate development of predictions of potential geographic range that are essential for 

effective control and eradication efforts. Current summer temperatures throughout most of the 

lower 48 states should support ALB survival and reproduction. Given ALB survival is brief 

where summer temperatures for a full day exceed 30o

C, continued future warming may 

adversely impact on the beetle’s vigor and establishment in specific regions of the country. 

Indirect impact of climate change on Asian longhorned beetle. The level of uncertainty 

regarding specific climate change impacts makes planning more challenging. The potential 

implications of this for ALB management cannot be understated. Forest-specific strategies for 

climate change adaptation are now urgently needed to inform management plans. Most cities and 

many natural forests across the northeast US and southeast Canada face threats 

from multiple insect pests; many of the most difficult to manage are exotic invasives that have 

no or few indigenous biological enemies. In addition to ALB, emerald ash borer (EAB), gypsy 

moth (GM) and Dutch elm disease (DED) pose significant threats and are being monitored 
D–12 Appendix D. Climate Change Impacts Related to ALB

closely as part of control programs to slow widespread damage (Nowak et al, 2010; Toronto, 

2007). Climate change is already having an impact by increasing the severity of outbreaks by 

these pests. This increases greatly the costs of their control, as well as raises the need for and 

expenditures on tree replacement and for programs designed to maintain tree health. A high 

premium is placed on strategic management of urban forests and pests to maintain and expand 

tree cover. Under warming conditions and enhanced needs for pest control efforts, there is 

increasingly less leeway for delay and discretionary management action. 

It is now clear that ALB and other exotic pests are having sizable impact on urban trees. For 

example, in Chicago, ALB could potentially cause losses to the urban forest of $1.3 billion in 

structural value (53.6% of live tree population), GM $595 million of losses (19% of all trees), 

EAB $295 million (12% of all trees) and DED, an additional $31 million (5.5 % of all trees). In 

Toronto, ALB discovered in 2003 can affect 43% of the City’s tree population and potentially 

cause structural losses estimated at $4 billion. Over the 1930-1950 period, DED killed as many 

as one-third to half of all City trees and could kill an additional 1.6% ($279 million); current 

outbreaks of GM and EAB threaten an additional 24% of the tree population ($2.1 billion 

structural value). One result (compounded by urban conversions) is that adequate tree cover in 

ALB-infested cities is difficult to maintain. Toronto’s tree cover diminished 7% over the 1985-

2005 period -- in spite of accelerated tree re-planting programs. Tree diversity is now less and 

vulnerability greater as individual species are selectively killed in successive pest outbreaks 

(Toronto 2007). In New Your City, ALB represents a potential loss to urban forest of $2.25 

billion in structural value (43.1% of the tree population); GM could potentially kill 23% of all 

trees ($2.21 billion loss); EAB 0.5% of all trees ($9.8 million loss), and DED an additional 0.7% 

of all trees ($111 million loss). 

A recommended tree cover target is an average of 40% canopy cover -- to ensure the 

sustainability of the urban forest and maximize the ecological, social and economic benefits 

derived from urban trees (American Forests 2007). Using this standard, Baltimore, Chicago, 

New York City and Toronto require increases of 16%, 23%, 20% and 20%, respectively, to 

attain the tree cover target (Nowak et al 2007, 2010, Toronto 2007). A tabulation of percent tree 

cover in 17 North American cities shows a range of 7% to 37%, with an average of about 25% 

(Nowak et al 2010) suggesting considerable room to expand urban tree cover to capture the 

benefits of CO2 sequestration, carbon storage and environmental benefits as climate changes. 

In summary, the nexus of climate change and the presence of ALB and other exotic pests is 

greatly increasing the demand for concerted management for tree health and sustainability of 

urban and adjacent natural forests. 
Appendix D. Climate Change Impacts Related to ALB D–13 

2.2.2 Direct and indirect impact of ALB eradication project on climate 

change

 

Direct impacts of eradication project on climate change. The most conspicuous impact of the 

removal of large numbers of ALB-infested trees is the loss of CO2 sequestration and release of 

carbon storage in trees and soil. Our estimates indicate that across the northeast US about 120,000 

trees have been removed and chipped, releasing a total of 20,000 mt of CO2 to the atmosphere 

since Year 2000. At risk over the five affected States is an estimated 2.4 million mt of CO2 

release from host trees (Table 1). 

 

If left to regenerate or if artificially replanted, the amount of CO2 released upon removal would 

be sequestered in new growth over an interval of 76 years on average (or less, if warming 

generally accelerates tree growth). A potential benefit is the removal of older infested trees. This 

assumes old trees more vulnerable than young age classes to disease and to damaging weather 

events (e.g., wind, snow, icing, drought), and are slower growing. 

Indirect impacts of eradication project on climate change. Tree removal under current ALB 

eradication programs serves to expand and exacerbate the impacts of climate change as we 

understand them. Below, we identify three categories of impacts where tree removal feeds back 

to exaggerate/exacerbate the on-going GHGs linked to global climate warming and attendant 

increases in extreme weather variations. 

 

(1) In urban areas, where trees now shade buildings, tree removal under ALB eradication 

will increase energy requirements and emissions of GHG from power plants -- to 

compensate for increased heating in winter and air conditioning in summer. 

 

Based on average state energy costs in February 2009, trees in Chicago are estimated 

to reduce energy costs from residential buildings by $360,000 annually. Trees are 

estimated to slightly increase the amount of carbon released by fossil-fuel based 

power plants. However, this estimated increase in emissions (1,200 tons) is more 

than offset by annual carbon sequestration by trees (25,200 tons) (Nowak et al 

2010). Trees in Toronto are estimated to reduce energy costs from residential 

buildings by $9.7 million annually. Trees also provide an additional $483,000 CND 

in value per year by reducing the amount of carbon released by fossil-fuel based 

power plants, representing a reduction of 17,000 mt of carbon emissions. These 

values could be increased through more strategic tree planting to maximize the 

potential energy effects of trees (Toronto, 2007). In New Your City, the energy and 

cost savings are even greater. Trees there are estimated to reduce energy costs from 

residential buildings by $11.2 million annually based on 2002 energy costs. Trees 

also provide an additional $167,000 in value per year by reducing the amount of 

carbon released by the City’s fossil-fuel based power plants (a net reduction of 

9,100 tons of carbon emissions) (Nowak et al., 2007). 

(2) Reduced tree cover typically results in loss of storm-water regulation and flood 

moderation. A single, average- size tree (21 m height) evapo-transpires 10-200 liters 

water per day or about 2000-40000 liters per year per tree, pending tree species and 

climate conditions (Wullschleger et al., 1997). Trees also lessen soil erosion and 
D–14 Appendix D. Climate Change Impacts Related to ALB

nutrient loss to streams and ground water. For example, of total annual 

precipitation, the combination of canopy interception, surface evaporation and 

transpiration loss is estimated to be 40-64% in deciduous broadleaf forests and 55-

80% in confers forests in the UK (Nisbet, 2005). Hence, trees exert a powerful 

modulating influence on water dynamics and this function is put at risk as tree 

cover is diminished. 

(3) By removing significant levels of air pollutants, urban trees and forests improve 

local air quality in five main ways (Nowak et al., 2007; 2010; Toronto, 2007), 

namely by: 

• absorbing gaseous pollutants through leaf surfaces, including ozone (O3) 

and nitrogen dioxide (NO2) 

• intercepting particulate matter (e.g., dust, ash, dirt, pollen, smoke) 

• reducing emissions from power generation for heating and cooling of buildings 

• releasing oxygen through photosynthesis 

• transpiring water and shading surfaces, lowing local air temperatures, and thereby 

reducing O3 levels. 

Not all gases above are GHGs but in total and across the composite of urban 

areas, these do affect local and regional climates (e.g., heat island effect) and 

ultimately feed-back variously to the global warming phenomenon. Although 

trees do emit volatile organic compounds (VOCs) that can contribute to ozone 

formation, integrative studies have revealed that an increase in tree cover actually 

leads to reduced ozone formation (Toronto, 2007). 

In Chicago trees remove an estimated 888 tons of air pollution (CO, NO2,O3, PM10, 

SO2) per year with an associated value of $6.4 million (Nowak et al., 2010). In 

Toronto trees and shrubs remove an estimated 1,430 mt of the same air pollutants per 

year, with an associated value of $16.1 million (Toronto 2007). In New York City, 

removal is estimated at 2,202 tons/year with an associated value of $10.6 million/year 

(Nowak et al., 2007). 

2.2.3 Manifestation of evident/known impacts of global climate 

change in geographic area

Keena and colleagues (Keena, 2006; 2009; Keena and Moore, 2010; Sanchez and Keena, 2013) 

demonstrate that climate has a strong impact on ALB growth and survival. Hence, the question -

- what actual evidence is available to demonstrate impacts on ALB populations have occurred 

already and continue to occur as a result of global climate change? 

Long-term shifts recorded for average temperature and total precipitation at Cincinnati, Ohio are 

both modest, highly variable, and may be cyclical in nature (Fig. 2a, 2b) (NCDC, 2013). Longterm air temperature records for southwestern Ohio (NOAA Zone 8) indicate an increase of 

0.42C / 100 years over the 1895 through 2013 period. This magnitude is modest relative to 

global temperature change, estimated to be 0.73C / 100 years (Attachment 3). 

 
Appendix D. Climate Change Impacts Related to ALB D–15 

Figure 2. Yearly (a) average air temperature and (b) average total precipitation in NOAA’s 

meteorological station at Cincinnati’s Lunken Field airport. Linear and 4th order 

polynomial regression trends and equations shown. Data obtained from 

http://www.esrl.noaa.gov/psd/data/timeseries/ (accessed 11/6/2013) includes observations 

of January 1895 through October 2013. 

 

A simple model of the ‘favorability’ of daily weather conditions, May through October over 

the 1950 – 2013 period was developed using Keena (2009) observations on ALB response to 

temperature and humidity conditions. The ALB Climatic Favorability Index is the sum of 

daily average temperatures 10 C to 33 C, minus the sum of temperature <10 C and > 32 C, 

multiplied by the sum of total precipitation > 0 mm, an calibrated on a scale of 0-100 by 

dividing by 10,000,000. 

The Index is dominated by a cyclic pattern. Based on the linear regression, there is a decrease of 

0.95 index points (or 1.6% per 100 years) relative to 1950. Since first detection on wood 

packaging material arriving from China in 1996, six years, (1996, 2000, 2001, 2003, 2004, 2009) 

stand out as particularly favorable for ALB growth (i.e., >17.5 Index); this contrasts to any earlier 

18-year period which had two to three such events over 18 years. Six years were relatively 
D–16 Appendix D. Climate Change Impacts Related to ALB

unfavorable (<7.5) over the 1996-2013 interval (Fig. 3). These Index values suggest an overall 

static trend but one punctuated by especially favorable or adverse yearly conditions. It is 

noteworthy that the present outbreak in Clermont County Ohio was preceded by a five year 

interval (2000-2004) of favorable climate for ALB, with the one exception of 2002. 

In brief, our analysis shows moderate temperature and precipitation shifts long-term in the 

southwest Ohio zone, consistent with but more moderate than global trends (Attachment 3). Our 

Index of climatic suitability for ALB shows both favorable and adverse years occur on a 

frequency of about 1.4 and 2.0 per decade, respectively, but with virtually no overall trend since 

1950. Cyclical patterns are conspicuous in all trends, implying a need for caution in stating 

definitively there is clear evidence of impact of global climate change on geographic area. 

Outbreak sites on the Great Lakes (Chicago, Toronto), the Atlantic Ocean (Worcester, Boston, 

New York City, Jersey City) and the Ohio River Valley (Cincinnati) imply that ALB thrives 

within a ‘humidity corridor’ and that commonly applied degree-day models require revision to 

fully incorporate moisture parameters. 

 

Figure 3. Daily estimate of favorability of climate for ALB survival and growth in 

Cincinnati, OH. ALB Climatic Favorability Index based on Keena et al. (Keena, 2006; 

2009; Keena and Moore, 2010; Sanchez and Keena, 2013) temperature and precipitation 

thresholds. Daily maximum and minimum air temperature and daily total precipitation 

January 1950 through October 2013 observed at Cincinnati Lunken FAA AP, OH, Station 

1916 (http://www.esrl.noaa .gov/psd/cgi-bin/data/timeseries/ accessed 11/6/2013). Dashed 

line is linear regression of the Index trend. 
Appendix D. Climate Change Impacts Related to ALB D–17 

3.0 Alternatives

3.1 No Action Alternative 

 

We posit the question of what would be the impact on CO2 release under hypothetical conditions 

in which there was no ALB eradication effort. Keena and Moore (2010) concluded that climatic 

conditions are suitable for ALB establishment and survival in all contiguous 48 States. To 

estimate release US-wide, we use tree carbon estimates for urban areas in 48 States published by 

Nowak and Crane (2000) and convert their weights of C to metric tons (mt) of CO2 using the 

standard 3.67 conversion factor. 

We then assume a percentage of all trees in any State are ALB host tree species. The percent of 

all trees that were ALB host species was the percent basal area of trees in each State that were 

known host trees estimated by Nowak et al. (2003). Second, we estimated the forest area affected 

adjacent to each urban area. The fraction of trees in forest areas at the last two ALB eradication 

sites is 2.33-fold greater than that removed from those two urban areas. We use this conversion 

(vs. that for all five sites) on the consideration that these are more representative of on-going and 

near-future conditions (compared to the ALB locations early in 2000). Third, the CO2 released 

from soil (as in the main CO2 model) was estimated as 25% of that released from the urban and 

adjacent forest trees removed. Total potential CO2 release is the sum of that released from urban 

trees, forest trees, and soils. 

This total was estimated for each of the 5 States with APHIS ALB eradication efforts to date. For 

comparison, we also estimate the total release for each of the 20 States in the Northeast plus 

North central regions, and for all 48 States (Attachment 4). 

Alternative 1a. In this alternative, only urban host trees and those in adjacent forests are 

involved. Estimates indicate that in the absence of ALB eradication there is a potential release of 

CO2 of 478 million mt in the five States, 2,111 million mt from the twenty Northeast plus North 

central States, and 2,255 million mt for all forty-eight States. These estimates represent potential 

releases under tree removal methods standardly used to date to eradication ALB found in and 

adjacent to city areas. 

Alternative 1b. Urban forest area is only 3.5% of US total area on average; even with the 

addition of adjacent forest areas (2.33 x that of the urban trees), the area involved is only 11.7% 

of total. Using the same hypothetical approach and assumptions as above but involving all host 

trees in the landscape (not just those in urban and adjacent areas), the estimated potential releases 

are two to six-fold larger (Attachment 5). Potential release of CO2 is estimated to be 1,030 

million mt in the five States, 6,959 million mt from the twenty Northeast plus North central 

States, and 13,906 million mt for all forty-eight States. 

3.2 Preferred Alternative

We posit several alternative CO2 management scenarios to achieve enhanced carbon storage and 

sequestration rates in trees. Two of these, increased tree health and expanded tree cover were 

identified as obvious opportunities in urban tree studies, notably those of Nowak et al (2007, 
D–18 Appendix D. Climate Change Impacts Related to ALB

2010) and the City of Toronto (Toronto 2007). The fact that soils typically contain significant 

quantities of carbon, their treatment during and after eradication efforts also merits attention. 

(1) Tree Health. Under programs of active urban tree and forest management, there is an 

opportunity for replacing trees removed and some of the current trees in poor or declining 

health with tree species resistant to ALB, and at the same time, better adapted to 

enhanced warming and climate variability. Beyond this are opportunities to increase tree 

protection from pests and improve growth of urban trees, thereby increasing CO2 

sequestration and storage. 

The fraction of urban trees in unhealthy condition is notable. For example, an urban forest 

health survey across 20 Northeast and North central States found only 23% of all trees were 

in good (22%) to excellent (1%) condition. The remaining 77% ranged from fair (27%), to 

poor (16%), or declining (12%) and otherwise improving and variable condition (18%, and 

4% ‘other’) (Pokorny, 1998) (http://www.na.fs.fed.us/spfo/pubs/uf/survey/execsum.htm

accessed 11 08 2013). 

(2) Tree Cover. Recent surveys over urban areas in the US (and Canada) show the percent 

in tree cover is a fraction of what the American Forests (2007) recommends as adequate. 

Compared to their standard of 40% tree cover, the average over 20 cities was 23.0%. The 

large five cities with active tree planting programs (Baltimore, Boston, Chicago, New York 

City, Toronto) had an average of 20.4% and could benefit by doubling their tree cover 

(Nowak et al., 2010). Re-planting can be difficult to achieve; many US cites over the past 

decade in fact show increased conversion to building, roads and other infrastructure, hence a 

relative decrease in urban tree cover (Nowak and Greenfield, 2012). 

(3) Soil Disturbance. Impacts of tree removal operations on soil surface disturbance 

are unreported in the ALB data and remain a large uncertainty in our GHG estimates. 

We posit two tree management scenarios designed to address the issues of sub-optimal tree 

health and sub-par total percent urban area in tree cover. Details of our procedure are 

documented in Attachment 8. 

In the first case, we run the CO2 Emissions and Sequestration Model to estimate the effect that a 

continuous increased tree growth rate of 10% would have on the number of years to 

compensation point (i.e., total sequestration in new growth = total emission from removed 

trees), and on the magnitude of additional CO2 sequestration. We consider ten percent is 

conservative considering that three-quarters of all trees are in fair, poor or worst condition. We 

assumed that better silviculture and protection against pests will be part of the effort to ensure 

the enhanced growth is sustained. 

The current Ohio eradication site was used to demonstrate what could be achieved in an actual 

operation. This alternative (Scenario 2) to a ‘business-as-usual’ approach (Scenario 1) shortened 

the compensation time by only two years but resulted in an additional 330 mt of CO2 capture 

(Fig. 4; Table 2). 
Appendix D. Climate Change Impacts Related to ALB D–19 

The Ohio eradication site was also used to estimate the additional CO2 sequestered in a 

hypothetical expansion of tree cover. We assumed an urban tree cover of 25% at the Ohio site 

and increased this 1.6-fold to attain the American Forests (2007) standard of 40% (Scenario 

while still ensuring the better silviculture that is part of Scenario 2. The compensation point was 

reached in 43 years -- or in about half the time of the ‘business as usual’ scenario (Scenario 1). 

The additional CO2 captured by year 85 at the peak of tree growth was significant. Fully1,471 mt 

more CO2 was sequestered and stored, or twice that in the ‘business as usual’ case (Fig. 4, Table 

2). 

Clearly, there is considerable discretionary leeway to improve the sequestration and capture of 

CO2 through active management of tree populations in the post-eradication period. Further gains 

can be made by reducing soil disturbance and by ensuring adequate or improved soil drainage 

properties. 

Figure 4. Estimates of CO2 sequestration in new tree growth and dynamic of CO2 net flux 

under alternative management options of the Ohio eradication site. Scenario 1 is business 

as usual. Scenario 2 is management for 10% increased tree growth rate with improved 

silviculture and pest protection. Scenario 3 is a 1.6-fold expansion of tree cover while 

maintaining enhanced tree growth and protection. Dashed blue line aids in visualizing 

timing of compensation points (total sequestration = total emissions) of each scenario. 
D–20 Appendix D. Climate Change Impacts Related to ALB

Table 2. Three alternative ALB post-eradication alternatives for management of tree CO2 

sequestration and other benefits. Scenario 1 is 'business-as-usual'. Scenario 2 is 

management for a continuous 10% acceleration of tree growth, while maintaining enhanced 

tree protection. Scenario 3 is urban tree cover expanded from 25% to 40% together with 

enhanced tree growth and protection. Years to achieve compensation, peak CO2 

sequestration level, and gain in total CO2 over 'business as usual' are given. 

 

4.0 Mitigation Measures

4.1 Reduction of GHG Emissions and Environmental Impacts

The President’s Council on Environmental Quality (CEQ) directs agencies to consider and 

evaluate reasonable alternatives to proposals (Title 40, Code of Federal Regulation, Part 1502.14 

(40 CFR 1502.14)). Alternatives proposed to address climate change issues need to be relevant to 

the proposed action’s purpose and need as well as technically and scientifically feasible. 

Alternatives may include mitigation measures to reduce GHG emissions, affect carbon cycling, 

or enhance adaptive capacity. Alternatives developed to respond to climate change issues should 

clearly relate to the cause-effect relationship between the proposal and climate change and have 

meaningfully different climate change-related effects when compared to the proposal and other 

alternatives. 

4.1.1 Project design to minimize GHG emissions and environmental 

impacts

ALB eradication approaches will continue to require removal of trees (and pesticide injections) 

as applied in current methods. The newly introduced regulations on climate change require 

attention to details of how to minimize release of carbon stored in trees, how to maximize CO2 

sequestration in the post-removal period, and enhance climate adaptation. A comparison of 

scenarios assessed in this study indicate large emissions in the unlikely event of no action to 

achieve eradication nation-wide. Alternatively, there is considerable management leeway to 

increase the rates of tree regeneration post-removal and expand tree cover to levels exceeding 

pre-treatment (Fig. 5). 
Appendix D. Climate Change Impacts Related to ALB D–21 

Figure 5. Summary of CO2 emissions under ‘No Action’ Alternative 1 and ‘Enhanced 

Action’ Alternative 2. (a) Alternative 1b is extreme case where all host trees in U.S. are 

killed; Alternative 1a is case where only urban and adjacent forest host trees are killed by 

ALB. (b) Alternative 2 is case where CO2 emissions and sequestration of ‘business-asusual’ (Scenario 1) is compared with 10% accelerated growth (Scenario 2), and accelerated 

growth plus expanded tree cover (Scenario 3) (see Table 2). 

accelerated 

growth and 

expanded 

tree cover

4.1.2 Operational changes

Two significant changes in ALB eradication efforts could significantly improve the efficacy of 

the ALB Eradication Project nationwide. The first is to enhance the surveillance efforts to find 

ALB early through improved surveillance technology. This could include models of especially 

high-risk years and locales through the development of GIS-based maps (e.g., Nowak et al., 

2003; Kalaris and Crane, 2013) coupled with routinely updated models of ALB climate 

suitability. Identified “hotspots” based on the most recent climate data could be the focus of 

intensive field detection surveys. Second, data collection during tree removal in the future could 

include the records of tree species, tree diameter, soil type and degree of soil disturbance 

following tree removal. These data would then be used in refining the current approach and 

model precision. 

4.1.3 Compensatory measures

The removal of infested and high-risk trees will result in emissions of CO2 (and possibly 

other GHG such as NOx) that feed into the global levels of GHG in the atmosphere. Our 

analyses show there are two main “offsets” that compensate for these initial emissions. 

First, there is recapture of emitted carbon through natural tree regeneration; beyond this, 

enhanced tree management options that can capture all and even more than the total CO2
D–22 Appendix D. Climate Change Impacts Related to ALB

released. Practical and economically feasible silviculture programs can be designed to both 

accelerate the rate of CO2 capture by regenerating trees and exceed initial loss levels. 

Second, tree cover expansion and continuing tree protection provide numerous advantages of 

climatic adaptation for urban areas, with measureable human benefits. Payoffs include reduced 

energy consumption, moderation and/or capture of air pollutants, improved storm- water 

management, and the presence of shade trees. Adding a strong diversity of tree species is 

important given successive waves of mortality by exotic pests in the past, and their continuing 

introduction and establishment (Aukema et al., 2010) 

In addition to the cities frequently referred to in this report (Chicago, New York City, Toronto), 

other cities in the Northeast have active tree expansion programs in progress, such as Boston 

(http://www. growbostongreener.org/gbg/ accessed 11/05/2013) and Baltimore 

(http://www.baltimoretreetrust.org/ accessed 11/05/2013). These urban programs provide insight 

into the practicality and cost advantages of on-going operations and the many added human 

benefits from increased tree cover in urban areas. 

5.0 References

Aber, J.D., S.V. Ollinger, C.A. Federer, P.B. Reich, M.L. Goulden, D.W. Kicklighter, J.M. 

Melillo and R.G. Lathrop. 1995. Predicting the effects of climate change on water yield and 

forest production in the northeastern United States. Clim. Res. 5: 205-222. 

http://cedarcreek.umn.edu/biblio/fulltext/t1002.pdf 

American Forests 2007. Urban tree cover forty percent standard. 

http://www.americanforests.org/ 

Arthur, M.A., L.M. Tritton, and T.J. Fahey. 1992. Dead bole mass and nutrients remaining 23 

years after clear-felling of a northern hardwood forest. Can J For Res 23:1298–1305. 

Aukema, J.E., D.G. McCullough, B. Von Holle, A.M. Liebhold, K. Britton and S.J. Frankel. 

2010. Historical accumulation of nonindigenous forest pests in the continental United States. 

BioScience 60: 886-897. 

Bebber, D.P., M.A.T. Ramotowski, and S.J. Gurr. 2013. Crop pests and pathogens move 

polewards in a warming world. Nature Climate Change, Advance Online Publication 

www.nature.com/natureclimate change doi: 10.1038/nclimate1990. 

CEQ. 2010. Draft NEPA Guidance on Consideration of the Effects of Climate Change and 

Greenhouse Gas Emissions. February 18, 2010 Release, Council on Environmental Quality. 

Washington DC 22 p. 

Christopher, S.F., M.J. Mitchell, M.R. McHale, E.W. Boyer, and D.A. Burns. 2008. Factors 

controlling nitrogen release from two forested catchments with contrasting hydrochemical 

responses. Hydrol. Process. 22: 46–62. 
Appendix D. Climate Change Impacts Related to ALB D–23 

Dale, V.H. and S.C. Beyeler. 2001. Challenges in the development and use of ecological 

indicators. Ecological Indicators 1:3-10.. 

Iverson, L.R. and A.M. Prasad 1998. Predicting abundance of 80 tree species following climate 

change in the eastern United States. Ecological Monographs 68:465-485. 

Kalaris, T. and M. Crane 2013. Asian Longhorned Beetle degree day modeling and usage. 

Geospatial Information System (GIS) Presentation, USDA APHIS, Riverdale MD November 19, 

2013. 

Karl, T.R., J.M. Melillo, and T.C. Peterson. 2009. Global Climate Change Impacts in the United 

States (eds.). Cambridge Univ. Press. 

http://downloads.globalchange.gov/usimpacts/pdfs/climate-impacts-report.pdf. 

Keena, M.A. 2006. Effects of temperature on Anoplophora glabripennis (Coleoptera: 

Cerambycidae) adult survival, reproduction, and egg hatch. Environ. Entomol. 35(4): 912-921. 

Keena, M.A. 2009. Phenology of the Asian longhorned beetle under simulated annual 

environmental fluctuations. In: K. McManus and K.W. Gottschalk eds. Proceedings, 19th U.S. 

Department of Agriculture interagency research forum on invasive species 2008: January 8-11 

2008; Annapolis, MD. Gen. Tech. Rep. NRS-P-36. Newtown Square, PA: U.S. Department of 

Agriculture, Forest Service, Northern Research Station: 37. http://www.nrs.fs.fed.us/pubs/gtr/gtrnrs-p-36papers/28keena-p-36.pdf 

Keena, M.A. and P.M. Moore 2010. Effects of temperature on Anoplophora glabripennis

(Coleoptera: Cerambycidae) larvae and pupae. Environ. Entomol. 39(4): 1323-1335. 

NCDC. 2013. National Climate Data Center. Cincinnati Municipal Airport Lunken Field, OH. 

US. http://www1.ncdc.noaa.gov/. 

Nisbet, T. 2005. Water use by trees. Information Note (April 2005) Forest Research, Forestry 

Commission, Surrey UK. 8 p. http://www.forestry.gov.uk/pdf/fcin065.pdf/$FILE/fcin065.pdf 

Nowak D.J. and D.E. Crane. 2002. Carbon storage and sequestration by urban trees in the USA. 

Environ. Pollution 116: 381–389. 

Nowak, D.J. and E.J. Greenfield. 2012. Tree and impervious cover change in U.S. cities. Urban 

Forestry and Urban Greening 11: 21– 30. 

Nowak, D.J., D. Crane, J. Stevens, and J. Walton 2003. Potential Damage from Asian 

Longhorned Beetle. Northern Research Station, USDA Forest Service. Syracuse, New York. 

(http://www.nrs.fs.fed.us /disturbance/invasive_species/alb/localresources/docs/ALB_potential.pdf accessed 11/13/2013) 
D–24 Appendix D. Climate Change Impacts Related to ALB

Nowak, D.J., R.E. Hoehn, D.E. Crane, J.C. Stevens, and C. Leblanc Fisher. 2010. Assessing 

urban forest effects and values, Chicago’s urban forest. Resour. Bull. NRS-37. Newtown Square, 

PA: U.S. Department of Agriculture, Forest Service, Northern Research Station. 27 p. 

http://www.itreetools.org/resources/ reports/ Chicago's%20Urban%20Forest.pdf 

Nowak, D.J., R.E. Hoehn, D.E. Crane, J.C. Stevens, and J.T. Walton 2007. Assessing urban 

forest effects and values, New York City’s urban forest. Resour. Bull. NRS-9. Newtown Square, 

PA: U.S. Department of Agriculture, Forest Service, Northern Research Station. 22 p. 

http://www.milliontreesnyc.org/ download /pdf/ufore_study.pdf 

Overpeck, J.T., P.J. Bartlein, and T. Webb III. 1991. Potential magnitude of future vegetation 

change in eastern North America: comparisons with the past. Science 254: 692-695. 

Pokorny, J.D. 1998. Urban forest health: Needs assessment survey, Results and 

recommendations. Report NA-TP-01-98 Northeastern State and Private Forestry, USDA Forest 

Service, St. Paul MN. 

Pouyat, R.V., M.J. McDonnell, and S. T. Pickett. 1997. Litter decomposition and nitrogen 

mineralization in oak stands along an urban-rural land use gradient. Urban Ecosystems 1:117– 

131. 

Sanchez, S.A. and M.A. Keena 2013. Development of the teneral Adult Anoplophora 

glabripennis (Coleoptera: Cerambycidae): time to initiate and completely bore out of maple 

wood. Environmental Entomology 42: 1-6. 

Tubby, K.V. and J.F. Webber. 2010. Pests and diseases threatening urban trees under a changing 

climate. Forestry 83: 451-459. 

Toronto. 2007. Every tree counts: a portrait of Toronto’s urban forest. (Forward by R. Ubbens, 

director), Parks, Forestry and Recreation: Urban Forestry. City of Toronto, Toronto, Ontario, 

Canada. 63 p. http://www.sfu.ca /grow/science/resources/1304545703.pdf 

USDA APHIS 2013. Asian Longhorned Beetle Cooperative Eradication Program in Clermont 

County, Ohio Eradication Program in Clermont County, Ohio. United States Department of 

Agriculture, Marketing and Regulatory Programs, Animal and Plant Health Inspection Service. 

Revised Environmental Assessment, Draft January 2013. (Agency Contact: Robyn Rose, 

National Program Manager Asian Longhorned Beetle Eradication Program, USDA APHIS PPQ, 

4700 River Road, Riverdale, MD 20737 107 p. 

USDA OCE. 2013. Science-based methods for entity-scale quantification of greenhouse gas 

sources and sinks from agriculture and forestry practices. Draft report Climate Change Program 

Office, Office of the Chief Economist, USDA. Washington DC http://usda.gov/oce/climate 

_change/estimation.htm 

Wilmont, S. 2011. Climate change and Vermont’s forests. Climate Change Adaptation White 

Paper Series (May 2011), Vermont Agency of Natural Resources, Secretary's Office, Montpelier, 

Vermont. http://www.anr.state.vt.us/anr/ climatechange/Pubs/VTCCAdaptForestry.pdf 
Appendix D. Climate Change Impacts Related to ALB D–25 

Wullschleger, S.D., F.C. Meinzer, and R.A. Vertessy. 1997. A review of whole-plant water use 

studies in trees. Tree Physiology 18: 499-512 

Zhou, J., K. Zhang, and Y. Lu. 1984. Study on adult activity and behavioral mechanism of 

Anoplophora nobilis Ganglbauer. Sci. Silv. Sin. 20: 379-379F. 
D–26 Appendix D. Climate Change Impacts Related to ALB

Attachments

Attachment 1

Major elements, sources of data and assumptions used in development of the CO2 Emissions and 

Sequestration Model. 

Trees Removed number of trees removed at each of five ALB Eradication Program sites 

Percent Urban percent of all trees in urban area and in forest area for each of five sites 

Tree Volumes total tree volume, including roots, trunks, branches, leaves estimated using 

tree species equations of USDA Forest Service COLE Program 

(http://www. ncasi2.org/COLE/), using ‘Individual Tree Spreadsheets’ 

option. Algorithm requires number of trees of selected tree species by 

diameter (dbh) class, tree age of dbh class, tree height of dbh class, and 

tree specific equation. Tree age and height estimates are the same as those 

of Nowak et al (2010). The example below is for sugar maple in urban 

areas: 

Tree Species Three tree species are selected to represent a range of coefficient b for tree 

volume representative of northern hardwoods, namely sugar or “hard” 
Appendix D. Climate Change Impacts Related to ALB D–27 

maple, red maple, and silver or “soft” maple. Tree species are not 

recorded for tree removals on eradication sites (Attachment 1). 

Urban and Forest Tree diameter (dbh) and tree heights differ between urban and forestgrown trees such that separate equations are used. 

Tree Dry Weights Tree volume converted to tree dry weight using coefficient b (above) of 

each tree species. Total dry weight of each tree component (e.g., foliage) 

for each dbh size class is summed to calculate total tree dry weight. For 

example, calculations for sugar maple in urban areas: 

Tree and Soil CO2 Conversion factors are used to estimate CO2 levels from estimates of dry 

weights. Nowak et al. (2010) use 0.8 to calibrate for urban trees that 

typically have less biomass than predicted by forest-derived biomass 

equations. The species carbon factor (0.5) is the carbon content of wood, 

converted to CO2 content using the standard 3.67coefficient. The total 

carbon at soil surface in urban areas is obtained from Pouyat et al. (1997). 

Soil carbon exposed is a fraction of the crown leaf area (above); this is 

multiplied by an assumed exposure factor (5%) likely once the tree canopy 

is removed. The total soil CO2 release is product of the three prior terms. 

Total CO2 is the sum of the tree and soil carbon. All US units are then 

standardized in metric tons (mt). 
D–28 Appendix D. Climate Change Impacts Related to ALB

CO2 Emission The rate of decomposition and release of CO2 to atmosphere from chipped 

wood following tree removal are based on those for deciduous hardwood tree species studied in 

the Northeast US (e.g., Arthur et al., 1992 at Hubbard Brook NH). Assumption is made that 

chipping wood of large boles does not significantly alter natural decomposition rate. Example of 

CO2 emission from decomposition of bole wood is shown below. 
Appendix D. Climate Change Impacts Related to ALB D–29 

Attachment 2

Coordinates of Cincinnati, OH US Meteorological Station, website and types of data accessed 

used in models of ALB Climatic Favorability Index for Alternative 2 evaluations. 

Latitude: 39° 14' 27" N (deg min sec), 39.2408° (decimal), 3914.45N (LORAN) 

Longitude: 84° 29' 58" W (deg min sec), -84.4995° (decimal), 08429.97W (LORAN) 

Elevation: 245 meters (804 feet) validated against 244 meters (799 feet) from NED 

Contiguous US 1/3E arc second elevation data 

Location: Cincinnati, OH US 

County: Hamilton, OH 

GHCND: USW00093812 Cincinnati Municipal Airport Lunken Field, Oh US 

Website: http://www1.ncdc.noaa.gov/

Data Types: TMAX - Maximum temperature (tenths of degrees C) 

TMIN - Minimum temperature (tenths of degrees C) 

PRCP - Precipitation (tenths of mm) 

Cincinn
D–30 Appendix D. Climate Change Impacts Related to ALB

Attachment 3

Comparison of trend in global mean temperature, 1880-2013, (http://data.giss.nasa.gov

/gistemp/tabledata/GLB.Ts.txt accessed 11/13/2013) with that of southwest Ohio, Zone 8, 1895- 

2013 (http://data.giss.nasa.gov/gistemp/tabledata/GLB.Ts.txt accessed 11/6/2013). 

Dark trend lines are 11-point moving averages. 
Appendix D. Climate Change Impacts Related to ALB D–31 

Attachment 4

Estimate of CO2 release across US 48 contiguous States under Alternative 1 (no action). 

Assumes full mortality of host tree species in urban and adjacent forests. Metric tons (mt) 

of carbon of urban areas in each State from Nowak and Crane (2002). Conversion to tons 

CO2 based on 3.67 carbon: carbon dioxide ratio. CO2 released in adjacent forest areas 

based on a multiplier of 2.33 (see text for details). CO2 released from soils is 25% of total 

CO2 release from urban plus forest trees. Estimates are total tree plus soil CO2 release in 

each 48 State, in 20 Northeast and Northcentral States (USDA Forest Service Region 9) 

and in the 5 States to date with ALB Eradication Programs. 
D–32 Appendix D. Climate Change Impacts Related to ALB

Attachment 5

Extreme estimate of CO2 release from mortality of all host tree species across US 48 

contiguous States under Alternative 1 (no action). Same as in Attachment 6 except that all 

host trees assumed to be killed (not just those in urban and adjacent forests). 
Appendix D. Climate Change Impacts Related to ALB D–33 

Attachment 6

Assumptions used in runs of the CO2 Emissions and Sequestration Model to estimate levels of 

sequestration under the three scenarios of Alternative 2. 

Scenario 1 All data and model algorithms as documented in Attachment 3 for ‘standard’ run 

of CO2 Emissions and Sequestration Model. 

Scenario 2 Initial total CO2 sequestration of Year 1 is compounded annually at rate of 0.10. 

This estimate of CO2 sequestration in accelerated growth is then added to total 

CO2 estimates of Scenario 1 standard run over full 100 year sequence post-tree 

removal by ALB Eradication Program. 

Scenario 3 Yearly estimates of total CO2 sequestration of Scenario 2 are multiplied by 15% 

increase over the extant 25% urban tree cover, or by 1.6. This provides an 

estimate of CO2 sequestration where total urban tree cover is 40% as 

recommended by standards for US cities (American Forests 2007). 
D–34 Appendix D. Climate Change Impacts Related to ALB

(This page is intentionally left blank.) 
Appendix E. Herbicide Environmental Risk Assessment E–1 

Appendix E: Herbicide Human Health and Ecological 

Risk Assessment: Asian Longhorned 

Beetle Eradication Program 
E–2 Appendix E. Herbicide Environmental Risk Assessment

Table of Contents

EXECUTIVE SUMMARY 

1.0 INTRODUCTION 

2.1 PROBLEM FORMULATION 

2.2 Chemical Description and Product Use 

2.3 Physical and Chemical Properties 

2.4 Environmental Fate 

2.5 Hazard Identification 

2.5.1 Mechanism of Action and Pharmacokinetics 

2.5.2 Acute Toxicity 

2.5.3 Sub-Chronic/Chronic Toxicity 

2.5.4 Carcinogenicity/Mutagenicity 

2.5.5 Development and Reproductive Effects 

2.5.6 Endocrine Effects 

2.5.7 Potential Additive, Antagonistic, or Synergistic Effects 

3.0. DOSE-RESPONSE ASSESSMENT 

3.1 Human Health Dose-Response Assessment 

3.1.1 Triclopyr 

3.1.2 TCP 

3.1.3 Imazapyr 

3.1.4 Metsulfuron-Methyl 

3.1.5 Drinking Water 

3.2 Ecological Effects Analysis 

3.2.1 Aquatic Effects Analysis 

3.2.2 Terrestrial Effects Analysis 

4.1 EXPOSURE ASSESSMENT 

4.2 Human Health Exposure Assessment 

4.2.1 Identification of Potentially Exposed Human Populations and Complete 

Exposure Pathways 

4.2.2 Exposure Evaluation 

4.3 Ecological Exposure Assessment 

4.3.1 Aquatic Exposure Assessment 

4.3.2 Terrestrial Exposure Assessment 

5.1 RISK CHARACTERIZATION 

5.2 Human Health 

5.3 Aquatic and Terrestrial Ecological Risk Characterization 

5.3.1 Aquatic 

5.3.2 Terrestrial Wildlife and Domestic Animals 

5.3.3 Terrestrial Invertebrates and Plants 
Appendix E. Herbicide Environmental Risk Assessment E–3 

6.0 UNCERTAINTIES AND CUMULATIVE IMPACTS 

7.0 REFERENCES 

ATTACHMENT A: Risk Estimations for Workers and General Public 

 
E–4 Appendix E. Herbicide Environmental Risk Assessment

EXECUTIVE SUMMARY

 

The United States Department of Agriculture (USDA) Animal and Plant Health Inspection 

Service (APHIS) proposes to use the herbicides triclopyr or as a mixture using triclopyr, 

imazapyr, and metsulfuron-methyl to control sprouting vegetation from stumps that are host 

species for the Asian longhorned beetle (ALB). These herbicide treatments are needed as a way 

to prevent reinfestation of ALB-host trees that have been removed as part of the ALB 

Eradication Program. The preferred method of control for stumps is physical removal; however, 

in some cases, the use of herbicides is required. 

USDA APHIS evaluated the potential human health and ecological risks from the proposed use 

of triclopyr, imazapyr, and metsulfuron-methyl for the ALB Eradication Program. The risks to 

human health are expected to be negligible based on limited exposure from the proposed use 

pattern of these herbicides (hand painting and backpack spraying). Exposure is greatest for 

workers who will apply the product. The potential exposure for workers is low with the proper 

use of required protective equipment. The potential exposure for the general public is also 

minimal. Risks are quantified for workers and the general public to represent extreme exposure 

scenarios including accidental conditions. The conservative risk evaluation results show that 

the hazard index for workers and the general public do not exceed one (the USEPA level of 

concern), indicating that the exposure is unlikely to cause adverse health effects. Therefore, 

triclopyr and triclopyr mixed with imazapyr and metsulfuron-methyl used in the ALB 

Eradication Program should pose minimal risk to human health. 

The risk of herbicide use to non-target fish and wildlife is also minimal. The proposed use 

pattern reduces potential exposure to most non-target fish and wildlife. Wild mammals and 

birds are at very low risk from herbicide applications due to the low toxicity of all three 

herbicides and the lack of anticipated effects to food sources that they use. Aquatic organisms 

are also at low risk based on the favorable toxicity profile for all three herbicides and expected 

residues that could occur in aquatic environments from the proposed applications. Non-target 

terrestrial plants are at the greatest risk from herbicide treatment; however, the method of 

application and selective use of herbicides as a treatment for stumps will reduce the risk to 

terrestrial plants. 

1.0 INTRODUCTION

This human health risk assessment (HHRA) and ecological risk assessment (ERA) provides a 

quantitative and qualitative evaluation of the potential risks and hazards to human health, 

nontarget fish, and wildlife from exposure to the herbicides triclopyr, imazapyr, and 

metsulfuron-methyl when used to control the regrowth of stumps of host species of the Asian 

longhorned beetle (ALB). 

The methods used in this HHRA to assess potential human health effects follow standard 

regulatory guidance and methodologies (NRC, 1983; USEPA, 2014a), and generally conform to 

other Federal agencies such as U.S. Environmental Protection Agency (USEPA), Office of 

Pesticide Programs (USEPA/OPP). The methods used in this ERA to assess potential ecological 

risk to nontarget fish and wildlife follow USEPA methodologies and other 
Appendix E. Herbicide Environmental Risk Assessment E–5 

published methodologies with an emphasis on those used by USEPA/OPP in the pesticide 

registration process. 

The risk assessment is divided into four sections. The first is the problem formulation 

(identifying hazard), followed by the the effects analysis or dose-response assessment, and then 

the exposure analysis (identifying potentially exposed populations and determining potential 

exposure pathways for these populations). The fourth section of the risk assessment integrates 

the information from the exposure and effects analysis to characterize the risks to human health 

and the environment. In addition, the uncertainties associated with the risk assessment and 

potential cumulative impacts are discussed. 

2.0 PROBLEM FORMULATION

APHIS is proposing the use of the herbicides triclopyr, or triclopyr mixed with imazapyr and 

metsulfuron-methyl to treat stumps and associated sprouts from host trees that were removed to 

prevent the further spread of ALB. When possible, APHIS will physically remove host trees 

along with the stumps to prevent reinfestation. However, physical removal of the stumps may 

not be possible in some situations such as when it is impractical to move stump removal 

equipment into an area, or there may be restrictions for habitat protection. In situations where 

stump removal is not feasible, APHIS will apply herbicides to treat the remaining stumps and 

associated sprouts. 

Triclopyr is an herbicide that was first registered in 1979 (USEPA, 1998; USDA FS, 2011a). 

Triclopyr imitates a plant hormone (indoleacetic acid) classified as an auxin, and is used to 

control woody plants and broadleaf weeds (Cox, 2000). Imazapyr is an herbicide first registered 

in 1985 (USEPA, 2006). Imazapyr is an imidazolinone herbicide that inhibits acetohydroxyacid 

synthase, an enzyme involved in the biosynthesis of amino acids such as leucine, isoleucine and 

valine (HSDB, 2014b). Imazapyr is a systemic, non-selective, pre- and post-emergent herbicide 

that acts by inhibiting the synthesis of branched-chain amino acids in plants (USEPA, 2006). It 

is used for the control of a broad range of terrestrial and aquatic weeds. Metsulfuron-methyl is a 

sulfonylurea herbicide that inhibits the enzyme that catalyzes the biosynthesis of branched-chain 

amino acids (valine, leucine, and isoleucine) which are essential for plant growth (USDA FS, 

2011c). 

APHIS proposes the use of two triclopyr formulations for the treatment of stumps, Garlon® 3A 

and Pathfinder® II. Garlon® 3A contains the active ingredient triclopyr triethylamine salt (TEA). 

Pathfinder® II contains the active ingredient triclopyr butoxyethyl ester (BEE). Pathfinder® II 

allows more flexibility in being able to treat the bark instead of direct application to cut areas of 

the stem. In addition, APHIS is proposing some foliar applications of Garlon® 3A that will be 

mixed with two other herbicides. The active ingredients in these two other herbicides are 

imazapyr and metsulfuron-methyl in Arsenal® and Escort® XP, respectively. They will be used 

to treat sprouting foliage from stumps that are removed as part of the eradication efforts. All 

applications will be made either by hand painting undiluted material on the stump or directly 

spraying stumps and/or sprouting foliage using a backpack sprayer. 
E–6 Appendix E. Herbicide Environmental Risk Assessment

The following sections discuss the chemical description and product use; physical and chemical 

properties; environmental fate; and hazard identification for these herbicides. 

2.1 Chemical Description and Product Use

Triclopyr or triclopyr acid (C7-H 4-Cl3-NO3) (CAS No. 55335-06-3) is the common name for 

[(3,5,6-trichloro-2-pyridinly)oxy]acetic acid. There are no active commercial products for 

triclopyr acid. Triethylamine (TEA) salt and butoxyethyl ester (BEE) are two commercial forms 

of triclopyr. Garlon® 3A (EPA Reg. No. 62719-37) contains the active ingredient TEA (44.4%) 

and Pathfinder® II (EPA Reg. No. 62719-176) contains the active ingredient BEE (13.6%). 

Garlon® 3A also includes the chelating agent ethylene diamine tetraacetic acid and ethanol 

(USDA FS, 2011a). The other ingredients of Pathfinder® II were not specified by the 

manufacturing company. 

The proposed application methods for triclopyr are painting the undiluted triclopyr formulation 

on the surface of stumps or directly spraying stumps and /or sprouting foliage using a backpack 

sprayer with triclopyr mixed with imazapyr and metsulfuron-methyl. In spray applications, the 

herbicide sprayer or container is carried by backpack. The nozzle on the wand or gun jet of the 

backpack sprayer should not be positioned higher than the handlers' waist, reducing the 

likelihood that the chemical will come into direct contact with the arms, hands, or face of the 

worker. In addition, a large coarse droplet size applied as close to the target area as possible is 

used to minimize the potential for drift. 

Garlon® 3A is used at rates of 3/4 to 9 pound (lb) acid equivalent (a.e.) of triclopyr (1/4 to 3 

gallons of Garlon® 3A) per acre to control broadleaf weeds and woody plants (Dow 

AgroSciences, 2011a). Pathfinder® II is used at no more than 8 lb a.e. per acre per year on noncrop areas (Dow AgroSciences, 2011b). The Garlon® 3A and Pathfinder® II label requirements 

for the restricted-entry interval are 48 hours and 12 hours, respectively (Dow AgroSciences 

2011a; 2011b). During the restricted entry interval, entry is not allowed except for workers with 

the proper personal protective equipment (PPE). 

Imazapyr (CAS No. 81334-34-1) is an imidazolinone herbicide while metsulfuron-methyl (CAS 

No. 74223-64-6) is a sulfonylurea herbicide. Both products are a common mix partners with 

triclopyr in the control of woody vegetation. Arsenal® (EPA Reg. No. 241-346) contains 27.8% 

or 28.7% isopropylamine salt of imazapyr and 72.2% or 71.3% other ingredients. The Escort® 

XP formulation (EPA Reg. No. 352-439) contains 60% metsulfuron-methyl and 40% other 

ingredients. For low-volume foliar brush control, the Arsenal®application rate is 0.5 to 1% by 

volume and 2 oz per acre when in a tank mix with Escort® XP. Garlon® 3A may be added to the 

mix at 1 to 2 pints per acre (BASF, 2012a). The Arsenal® and Escort® XP label requirements for 

the restricted-entry interval are 48 hours and 4 hours, respectively (BASF, 2012a; DuPont, 

2012a). 

2.2 Physical and Chemical Properties

Triclopyr is a colorless solid with a melting point of 148-150o

C. Triclopyr TEA is a grayish 

white granular solid with a melting point of 111-117o

C (USEPA, 1998). Triclopyr acid, 
Appendix E. Herbicide Environmental Risk Assessment E–7 

triclopyr TEA, and triclopyr BEE have vapor pressures of 1.26 x 10-6, <1 x 10-8, and 

3.6 x 10-6 mm Hg, respectively, indicating that these compounds can volatilize into vapor and be 

transported as a vapor or in aparticulate phase into the ambient air. Triclopyr acid and triclopyr 

TEA are soluble (water solubility of 430 mg/L and 4.12 x 105 mg/L, respectively) (HSDB, 

2014a). Triclopyr BEE has relatively low solubility (6.8 mg/L) (USEPA, 1998). The basic 

physical and chemical properties of the commercial products, triclopyr TEA and triclopyr BEE 

are summarized in table 2-1. 

 

Table 2-1. Physical and chemical properties for triclopyr TEA and BEE. 

Parameters Triclopyr TEA Triclopyr BEE

CAS No. 57213-69-1 64700-56-7 

Molecular Formula C13H19Cl3N2O3 C13H16Cl3NO4 

Molecular Structure 

Molecular Weight 358.67 356.63 

Melting Point (o

C) 111-117oC NA 

Henry Law Constant 

(atm m3 mol-1) 1.15 x 10-14 2.47 x 10-7

Vapor Pressure 

(mm Hg) (25o

C) <1 x 10-8 3.6 x 10-6

Water Solubility, 

mg/L 412,000 7.4 

Log Kow 1.5011 4.0133 

Sources: USEPA, 1998; HSDB, 2014a 

 Imazapyr is a clear, slightly viscous, pale yellow to dark green aqueous liquid, or white to tan 

powder with a slight ammonia odor. Its melting point is 171o

C. Imazapyr has a low vapor 

pressure (1.79x10-11 mm Hg), and low Henry's Law constant (7.08x10-17 atm-cu m/mol), 

suggesting low volatility from soil and water. It is considered highly soluble (water solubility of 

1.13x104 mg/L) (HSDB, 2014b). The basic physical and chemical properties of imazapyr are 

summarized in table 2-2. 

Metsulfuron-methyl is a white or colorless crystal, or white to pale yellow solid with a faint, 

sweet, ester-like odor with a melting point of 163oC. Metsulfuron-methyl has low vapor 

pressure (2.5x10-12 mm Hg), and low Henry's Law constant (1.32x10-16 atm-cu m/mol). It also 

has a high solublility in water (water solubility of 9.50x103 mg/L) (HSDB, 2014c). The basic 

physical and chemical properties of metsulfuron-methyl are summarized in table 2-2. 
E–8 Appendix E. Herbicide Environmental Risk Assessment

Table 2-2. Physical and chemical properties for imazapyr and metsulfuron-methyl. 

Parameters Imazapyr Metsulfuron-methyl

CAS No. 81334-34-1 74223-64-6 

Molecular Formula C13H15N3O3 C14H15N5O6S 

Molecular Structure 

Molecular Weight 261.28 381.37 

Melting Point (o

C) 171 163 

Henry Law Constant (atm

m3 mol-1) 7.08 x 10-17 1.32 x 10-16

Vapor Pressure

(mm Hg) (25oC) 9.0 x 10-11 2.5 x 10-12

Water Solubility, mg/L 1.13E+04 9.5E+03 

Log Kow 0.22 2.2 

Source: HSDB, 2014b, c 

 2.3 Environmental Fate

The environmental fate describes the processes by which triclopyr, imazapyr, and metsulfuronmethyl move and degrade in the environment. Environmental fate processes include: 1) 

mobility, persistence, and degradation in soil, 2) movement to air, 3) migration potential to 

groundwater and surface water, and 4) plant uptake. 

In soil, triclopyr is expected to have high mobility (Koc's ranging from 1.5 to 134 mL/g organic 

carbon (oc)). Under anaerobic conditions, triclopyr is persistent with a half-life of approximately 

1,300 days. Under aerobic conditions, triclopyr biodegrades in silty clay loam and silty loam 

soils with half-lives of 8 and 18 days, respectively. The half-life of the major metabolite, 3,5,6- 

trichloro-2-pyridinol (TCP), is 30 to 90 days (HSDB 2014a). In soil, both forms of triclopyr 

degrade into several intermediates before ultimately degrading to carbon dioxide (CO2) (NPIC, 

2002). In air (ambient atmosphere), triclopyr is expected to exist in both the vapor and 

particulate phase based on its vapor pressure (1.26 x 10-6 mm Hg at 25 o

C). Triclopyr in the 

vapor phase will degrade to hydroxyl radicals through a photochemical reaction with an 

estimated half-life of 3.3 days. Triclopyr undergoes photodecomposition with a half-life of <12 

hours. In water, triclopyr is mainly broken down by exposure to sunlight. The half-life of 

triclopyr in water ranges from 1 to 10 days depending on water conditions such as turbidity 

(NPIC, 2002). Triclopyr degrades slowly in a soil:water system incubated aerobically, with a 

reported half-life of 142 days (HSDB 2014a). The potential for bioconcentration in aquatic 

organisms is low with an estimated bioconcentration factor (BCF) of 3. Triclopyr’s half-life in 

plants ranges from 3 to 10 days, and the primary metabolite is 3,5,6-trichloro-2-methoxypyridine 

(NPIC, 2002). 
Appendix E. Herbicide Environmental Risk Assessment E–9 

In soil, imazapyr is expected to have high mobility (Koc’s ranging from 8.2 to 110 mL/g oc for 

soil) (USEPA, 2007). Imazapyr biodegrades with aerobic soil half-lives ranging from 17.7 to 

63.1 days. In ambient atmosphere, imazapyr will likely exist in the particulate phase based on its 

vapor pressure (an estimated vapor pressure of 9.0x10-11 mm Hg at 25o C). In water, the 

potential for bioconcentration in aquatic organisms is low due to an estimated BCF of 3. 

Exposure to ultraviolet (UV) light in aqueous solutions results in complete degradation of 

imazapyr in 48 hrs, with a half-life of 7 hrs (HSDB, 2014b). 

In soil, metsulfuron-methyl is expected to have moderate to high mobility (Koc values ranging 

from 4 to 345 mL/g oc). Metsulfuron-methyl is more mobile in alkaline soils than in acidic 

soils. A typical half-life for metsulfuron-methyl in soil is 30 days (ranging from 14 to 180 days). 

Soil temperature, moisture content, and pH influence degradation of metsulfuron-methyl. 

Metsulfuron-methyl degrades faster under high moisture content and high temperature and 

acidic conditions (Trevathan, 2002). In air, metsulfuron-methyl will exist solely in the 

particulate phase (vapor pressure of 2.50x10-12 Hg at 25oC). Metsulfuron-methyl may undergo 

direct photolysis based on 50 and 76% degradation in an aqueous solution after a 15 and 36 

hour exposure to UV irradiation, respectively. In water, metsulfuron-methyl may undergo 

direct photolysis. It is not expected to adsorb to suspended solids and sediment based upon its 

Koc values. Metsulfuron-methyl has BCF values ranging from 1-17 suggesting the potential for 

bioconcentration in aquatic organisms is low (HSDB, 2014b). 

2.4 Hazard Identification

Technical triclopyr acid, TEA, and BEE are slightly toxic for the oral and dermal exposure 

routes (Toxicity Category III). They are practically non-toxic for the inhalation exposure route 

(Toxicity Category IV) and are not dermal irritants. However, triclopyr TEA was corrosive 

and BEE was minimally irritating in the primary eye irritation study. Triclopyr TEA and BEE 

are dermal sensitizers. The primary target organs of triclopyr are the liver and kidney 

(USEPA, 1998; USEPA, 2002a). 

Imazapyr is practically non-toxic for oral (Toxicity Category IV) and dermal (Toxicity Category 

III) routes of exposure. For the acute inhalation route of exposure, imazapyr is classified in 

Toxicity Category II. It is not irritating to the skin, and is negative for dermal sensitization. 

However, imazapyr causes acute irreversible eye damage (Toxicity Category I) (USEPA, 2006). 

Metsulfuron-methyl is non-toxic for oral, dermal, and inhalation exposure routes (Toxicity 

Category IV). It is not a skin irritant or sensitizer, but is a slight eye irritant (USEPA, 2002b). 

2.4.1 Mechanism of Action and Pharmacokinetics

Triclopyr, a weak acid, is excreted primarily from the kidney through an active transport 

process (Timchalk and Nolan, 1997; Timchalk et al., 1990; 1997). At very high doses, 

it may interfere with the excretion of other weak acids. However, concentrations of weak 

acids in the body under normal environmental exposures will be far below levels that 

environmental exposures will be far below levels that would interfere in the active transport 
E–10 Appendix E. Herbicide Environmental Risk Assessment

process. Therefore, this mechanism of active transport is not would i n e in the active transport 

process. Therefore, this mechanism of active transport is not expected to play a substantial role 

in potential health effects. Imazapyr is a plant amino acid synthesis inhibitor, which inhibits 

acetolactate synthase (ALS), an enzyme found only in plants and microorganisms. Plants 

require ALS for the synthesis of essential branched chain amino acids for their growth. 

Animals lack ALS and do not synthesize these amino acids by themselves (USDA FS, 2011b). 

A mechanism of action for imazapyr in mammals is currently unknown (USEPA, 2005a; 

HSDB, 2014b). Metsulfuron-methyl is also a plant amino acid synthesis inhibitor with the 

same mechanism of action as imazapyr. The mechanism of action of metsulfuron-methyl in 

mammals is not clear (USDA FS, 2005). 

Pharmacokinetic studies show that triclopyr is absorbed and excreted almost exclusively in the 

urine through acid hydrolysis (Shackelford et al., 1999). Following oral exposure, triclopyr is 

absorbed and excreted relatively rapidly, with half-lives for oral absorption and urinary 

excretion of 3.61 and 1.1 hours, respectively. A majority of ingested triclopyr is excreted 

unchanged in the urine, although minor metabolites can be formed (USFS, 2011a). Triclopyr 

has very low potential to be absorbed through the skin or to accumulate in humans at acutely 

toxic levels because it is absorbed slowly through the skin and is rapidly eliminated (HSDB, 

2014a). Imazapyr and metsulfuron-methyl are absorbed and excreted through urine and feces 

mostly unchanged (HSDB 2014b,c). In a rat study, 87% of an orally administered imazapyr 

dose was excreted in the urine and feces within 24 hours (HSDB, 2014b). Metsulfuron-methyl 

is eliminated from rats quickly (9-16 hours at low doses, and 23-29 hours for high doses) 

(HSDB, 2014c). 

2.4.2 Acute Toxicity

The acute oral rat LD50 values for TEA, Garlon® 3A, BEE and Pathfinder® II formulations show 

low toxicity (Category III) (table 2-3). Acute toxicity values for BEE are lower compared to 

those for TEA and Garlon® 3A, but are still within the same toxicity category. The TEA, BEE 

and Pathfinder® II formulations have low dermal toxicity (Category III) and Garlon® 3A has 

very low (Category IV) acute dermal toxicity. The acute inhalation toxicities for TEA, Garlon® 

3A, BEE and Pathfinder® II formulations are all very low (Category IV). None of the 

formulations are dermal irritants. Triclopyr TEA, Garlon® 3A, and BEE are dermal sensitizers. 

However, Pathfinder® II is not a dermal sensitizer. The primary eye irritation study results 

between the technical active ingredients and the proposed formulations vary from corrosive 

(triclopyr TEA - Category I), to may cause severe irritation with corneal injury (Garlon® 3A), 

and from minimally irritating (BEE) to slight temporary eye irritation (Pathfinder®II). 
Appendix E. Herbicide Environmental Risk Assessment E–11 

Table 2-3. Comparative acute mammal toxicity between triclopyr technical active 

ingredients and proposed formulations. 

 

Sources: USEPA, 1998; Dow AgroSciences, 2010; 2011c 

 Acute mammal toxicity data for imazapyr and Arsenal® show very low toxicity for oral, 

dermal, and inhalation routes (table 2-4). The formulation is much less toxic causing no eye 

irritation while the technical imazapyr can cause irreversible eye damage. 

 

Table 2-4. Comparative acute mammal toxicity between imazapyr technical active ingredient 

and the proposed formulation. 

Toxicity Study Imazapyr Technical Arsenal®

Acute Oral LD50 (rat) >5,000 mg/kg (IV) >5,000 mg/kg (IV)

Acute Dermal LD50 (rat) >2,000 mg/kg (III) >2,000 mg/kg (III)

Acute Inhalation LC50 (rat) 1.3 mg/L(gravimetric)

5.1 mg/L (nominal) (III) 

>5.3 mg/L (IV)

Primary Eye Irritation (rabbit) Irreversible Eye Damage (I) Non-irritating (IV) 

Primary Skin Irritation Non-irritating to slight erythema and 

edema (IV) 

Mildly-irritating (IV) 

Dermal Sensitization (Guinea pig) Negative Negative 

 Sources: USEPA, 2005a; BASF, 2012b

Toxicity 

Study 

TEA 

(44.4% 

a.i.)

Garlon® 3A BEE 

(97.1% a.i.)

Pathfinder® II 

Acute Oral 

LD50 (rat) 

1,847 

mg/kg 

(M&F) 

(III) 

1,847 mg/kg (III) 803 mg/kg (M&F) (III) 1,000 mg/kg 

(F) (III) 

Acute Dermal 

LD50 (rat for 

technical and 

rabbit for 

Garlon® 3A) 

>2,000 

mg/kg 

(III) 

>5,000 mg/kg (IV) >2,000 mg/kg (III) >2,000 mg/kg 

(III) 

Acute 

Inhalation 

LC50 (rat) 

>2.6 

mg/L 

(IV) 

>2.6 mg/L (IV) >4.8 mg/L (IV) >5.0 mg/L 

(IV) 

Primary Eye 

Irritation 

(rabbit) 

Corrosive 

(I) 

May cause severe 

irritation with corneal 

injury 

Minimally irritating (III) Slightly 

temporary 

irritation 

Primary Skin 

Irritation 

Not 

irritating 

(IV) 

Brief contact- not 

irritating; prolonged 

contact may cause 

slight skin irritation 

with local redness; 

repeated contact may 

cause skin burns 

Not irritating (IV) Prolonged skin 

contact may 

cause 

moderate skin 

irritation with 

local redness 

Dermal 

Sensitization 

(Guinea pig) 

Sensitizer Sensitizer Sensitizer Non-sensitizer 

 
E–12 Appendix E. Herbicide Environmental Risk Assessment

Acute mammal toxicity data for imazapyr and Arsenal® show very low toxicity for oral, 

dermal, and inhalation routes (table 2-4). The formulation is much less toxic causing no eye 

irritation while the technical imazapyr can cause irreversible eye damage. 

Acute mammal toxicity for technical metsulfuron-methyl and Escort® XP show low to very 

low toxicity for oral, dermal, and inhalation routes (table 2-5). The formulation is less toxic for 

the dermal route and causes less eye and skin irritation. 

Table 2-5. Comparative acute mammal toxicity between metsulfuron-methyl technical active 

ingredient and proposed formulation. 

Toxicity Study Metsulfuron-methyl Technical Escort

® XP

Acute Oral LD50 (rat) >5,000 mg/kg (IV) >5,000 mg/kg

Acute Dermal LD50 (rat) 3,000 mg/kg (III) >5,000 mg/kg

Acute Inhalation LC50 (rat) >2 mg/L (IV) >5 mg/L 

Primary Eye Irritation (rabbit) Irritation reversible in 7 days Slight irritation

Primary Skin Irritation Moderate irritation in 72 hours No irritation

Dermal Sensitization (Guinea pig) Negative Negative 

Sources: USFS, 2005; DuPont, 2012b 

 

2.4.3 Sub-Chronic/Chronic Toxicity

Based on chronic and subchronic toxicity studies, the kidney appears to be the most sensitive 

target organ for triclopyr. Decreased phenolsulfonphthalein urinary excretion and reduced 

absolute and relative kidney weights in dogs were observed after exposure to a dose of 

2.5 mg/kg/day for 183/184 (male/female) days (USDA FS, 2011a). Kidney effects on 

hematological and histopathological changes and increased kidney weight in rodents were 

observed after subchronic exposure to triclopyr doses at 70 mg/kg/day for 90 days. Damage was 

characterized as degeneration of the proximal tubules of the kidneys (≥20 mg/kg/day for 90 

days) and increases in kidney weight (USDA FS, 2011a). The No Observed Adverse Effect 

Level (NOAEL) for kidney toxicity in rats is 5 mg/kg bw/day from a two generation dietary 

reproduction study in rats. This is the basis of the chronic reference dose (RfD) for triclopyr 

(USEPA, 2002a). Other general systemic effects of triclopyr include signs of liver damage and 

a decrease in food consumption, growth rate, and gross body weight occurring at high doses 

(USDA FS, 2011a). 

Subchronic and chronic studies using imazapyr have been conducted in rats, rabbits, and dogs. 

The 90-day oral toxicity study in rats shows that the dermal and systemic NOAEL was 

1,695 mg/kg/day for males and 1,784 mg/kg/day for females at the highest dose tested (HDT). 

The 21/28-day dermal toxicity study in rabbits showed that the dermal and systemic NOAEL 

was 400 mg/kg/day (HDT). The chronic toxicity study in dogs (1 year) showed that the 

NOAEL was 250 mg/kg/day (HDT). The NOAEL of 250 mg/kg/day was selected as the basis 

of the chronic RfD for imazapyr because it was the lowest NOAEL. The dose of 250 

mg/kg/day was the highest dose tested in the dog study and no adverse effects were observed. 

USEPA used this dose for calculating the chronic RfD based on a structural analog of imazapic 

to choose a toxic endpoint. Imazapic causes skeletal muscle effects in dogs at 137 mg/kg/day 

(male) and 180 mg/kg/day (female) (USEPA, 2006). 
Appendix E. Herbicide Environmental Risk Assessment E–13 

Subchronic and chronic studies with metsulfuron-methyl have been conducted using rats, 

rabbits, and dogs. The 90-day oral toxicity study in rats reported a NOAEL of 68 mg/kg/day 

(male) and 64 mg/kg/day (female) with a Lowest Observable Adverse Effect Level (LOAEL)of 

521 mg/kg/day (male) and 659 mg/kg/day (female) based on a transient decrease in body 

weight. The 21-day dermal toxicity study in rabbits showed that the dermal NOAEL was 125 

mg/kg/day with a LOAEL of 500 mg/kg/day based on skin lesions (diffuse/multifocal 

dermatitis). The systemic NOAEL was 125 mg/kg/day with a LOAEL of 500 mg/kg/day based 

on an increased incidence of diarrhea. The chronic toxicity study in male and female rats 

showed that the NOAEL was 25 mg/kg/day with a LOAEL of 250 mg/kg/day based on body 

weight loss. The NOAEL of 25 mg/kg/day was selected as the basis for the chronic RfD for 

metsulfuron-methyl because it was the lowest NOAEL observed in the toxicity studies 

(USEPA, 2002b). 

2.4.4 Carcinogenicity/Mutagenicity

Based on the USEPA/OPP chemical evaluation of carcinogen potential (USEPA, 2013), 

triclopyr is “Not Classifiable as to Human Carcinogenicity (Group D)”. Triclopyr was 

classified as a Group D chemical because the evidence of the increase in mammary tumors in 

the female rat and mouse, and adrenal pheochromocytomas in the male rat are marginal. 

There was no additional support from structural analogs or genotoxicity studies (USEPA, 

1998). Triclopyr is not considered mutagenic or genotoxic based on a lack of evidence from 

several in vitro and in vivo studies (USEPA, 2002a). However, two unpublished studies on 

triclopyr ingestion by rats and mice have suggested increased frequency of mammary gland 

cancer at high doses. Mutagenicity studies using triclopyr (triclopyr technical acid, triclopyr 

BEE, and non specified triclopyr) in mice and rat did not show mutagenicity (USDA FS, 

2011a). 

USEPA (2013a) classifies imazapyr as “Evidence of Noncarcinogenicity for Humans (Group 

E)” based on a study in mice that showed no evidence of carcinogenicity. In the carcinogencity 

study, the NOAEL was 10,000 ppm (1,301 mg/kg/day in males and 1,639 mg/kg/day in 

females). The level was the HDT and there was no LOAEL (USEPA, 2005a). 

Imazapyr is not considered mutagenic or genotoxic based on a lack of evidence from 

several in vitro and in vivo studies (USEPA, 2005a). The results for the Ames assay was 

negative up to 5,000 μg/plate. The results of in vitro mammalian cell gene mutation and 

chromosome aberration studies were negative up to subchronic toxic doses (5,000 μg/ml) with 

and without activation. 

USEPA (2013a) classifies metsulfuron-methyl as “Not Likely to Be Carcinogenic to Humans” 

based on studies in rats that show no evidence of carcinogenicity. In one carcinogenicity study, 

the NOAEL was 25 mg/kg/day for both male and female rats. The LOAEL was 250 mg/kg/day 

for both male and female rats based on reduced body weight. In another carcinogenicity study, 

the NOAEL was 666 mg/kg/day (male) and 836 mg/kg/day (female) (USEPA, 2002b). 

Metsulfuron-methyl is not considered 
E–14 Appendix E. Herbicide Environmental Risk Assessment

mutagenic or genotoxic based on a lack of evidence from several in vitro and in vivo studies 

(USEPA, 2002b). 

2.4.5 Development and Reproductive Effects

The current chronic RfD for triclopyr was based on a two-generation reproduction study 

(USEPA, 1998). In this study, male and female rats were exposed to dietary concentrations of 

triclopyr resulting in doses of 0, 5, 25, or 250 mg/kg/day, except that the parent males in the 

high dose group were exposed to a concentration resulting in a daily dose of 100 mg/kg 

bw/day. The 5 mg/kg/day dose groups showed no evidence of adverse effects in the parents 

or offspring. At 25 mg/kg/day, degeneration of renal proximal tubules was observed only in 

adult animals. At 250 mg/kg/day, parental effects included decreased food consumption and 

body weights as well as histopathological changes in the liver and kidney. Fetotoxic effects, 

including decreased pup survival and litter sizes, were noted at 250 mg/kg/day. The NOAEL 

of 25 mg/kg/day for reproductive effects from this study is supported by a three-generation 

reproduction study using the same strain of rats with no adverse effects observed to offspring 

at doses of 3, 10, or 30 mg/kg/day (Hanley et al., 1984); and by an earlier study summarized 

by the Forest Service in which no adverse reproductive effects were observed in rats exposed 

to doses up to 30 mg/kg bw/day (USDA FS, 2011a). 

The developmental studies on triclopyr acid, triclopyr TEA, and triclopyr BEE show that 

triclopyr can cause adverse developmental effects including birth defects at sufficiently high 

doses. The developmental studies for the triclopyr salt and ester in rats show that the maternal 

NOAEL was 100 mg/kg/day, with a LOAEL of 300 mg/kg/day based on mortality (triclopyr 

salt and ester), clinical signs, necropsy findings, decreased body weight gains, decreased food 

consumption, increased water consumption, and increased relative kidney and liver weight 

(triclopyr ester). The developmental NOAEL was 100 mg/kg/day with a LOAEL of 300 

mg/kg/day based on decreased fetal weight, increased fetal and litter incidence of skeletal 

anomalies, increased fetal incidence of unossified sternebrae (triclopyr salt), increased 

incidence of hydrocephalus, cleft palate, microphthalmia/anophthalmia, retinal folds, thin 

diaphragm/protrusion of the liver, decreased fetal weight, and visceral and skeletal amomalies 

and variants (triclopyr ester). The maternal and developmental NOAELs and LOAELs in 

rabbits were 30 mg/kg/day and 100 mg/kg/day, respectively (USEPA, 2002a). A consistent 

pattern with triclopyr, however, is that adverse developmental effects occur only at doses that 

are maternally toxic. The developmental studies conducted in rats and rabbits do not suggest 

substantial or consistent differences in the developmental effects of the various forms of 

triclopyr (USDA FS, 2011a). 

The developmental studies in rats on imazapyr show that the maternal NOAEL was 300 mg/kg 

bw/day with a LOAEL based on salivation at 1,000 mg/kg bw/day. The developmental 

NOAEL was 1,000 mg/kg/day (HDT) and no LOAEL was reported. The developmental studies 

in rabbit showed that the maternal NOAEL was 400 mg/kg bw/day (HDT) and the 

developmental NOAEL was 400 mg/kg bw/day (HDT). The reproduction studies in rats 

showed that the NOAEL for parental systemic, reproductive andoffspring was 

10,000 ppm (738 mg/kg bw/day in males and 933.3 mg/kg bw/day in females) (HDT) (USEPA, 

2005a). 
Appendix E. Herbicide Environmental Risk Assessment E–15 

The developmental studies in rodents on metsulfuron-methyl show that the maternal NOAEL 

and LOAEL were 250 mg/kg/day, and 1,000 mg/kg/day, respectively, with effects on salivation 

and decreased body weight gain-compensatory increase after dosing stopped. The 

developmental NOAEL and LOAEL were <1,000 mg/kg/day (HDT) and >1,000 mg/kg/day 

(HDT), respectively. The developmental studies in nonrodents show that the maternal NOAEL 

and LOAEL were 25 mg/kg/day, and 1,000 mg/kg/day, respectively, with effects on increased 

mortality, decreased body weight gains, and clinical signs of anorexia, red/orange urine and/or 

exudate. The developmental NOAELwas 700 mg/kg/day (HDT). Reproduction studies in rats 

showed that the NOAEL for parental systemic was 34 mg/kg/day and 43 mg/kg/day (male and 

female respectively). The LOAEL was 342 and 475 mg/kg/day (male and female 

respectively) based on decreased premating body weight gains. The reproductive NOAEL was 

342 and 475 mg/kg/day (male and female respectively) (HDT). The offspring NOAEL was 

342 and 475 mg/kg/day (male and female respectively) (HDT) (USEPA, 2002b). 

2.4.6 Endocrine Effects

A literature search did not identify any study indicating the potential for triclopyr, imazapyr, 

and metsulfuron-methyl to disrupt endocrine function. None of these herbercides are among 

the group of pesticide active ingredients to be screened under the USEPA Endocrine Disruptor 

Screening Program. However, the identity of chemicals for screening is based on exposure 

potential, not on whether the pesticide is a known or likely potential endocrine disruptor 

(USEPA, 2014b). Fetal toxicity and abnormalities have been observed at higher doses using 

triclopyr, however, there is no indication that the effects occurred through a mechanism 

involving endocrine disruption (USDA FS, 2011a). 

2.4.7 Potential Additive, Antagonistic, or Synergistic Effects:

A literature search was performed to identify any studies indicating additive (toxicity of the 

mixture is equivalent to the sum of the toxicities of the individual compounds), antagonistic 

(toxicity of the mixture is less than additive), or synergistic (toxicity of the mixture is greater 

than additive) effects from the mixing of the three herbicides. The search results indicate that 

the most common type of effect is additive toxicity and there may be mildly synergistic effects 

associated with the proposed herbicide mixture (Tatum, 2004). The formulations of the three 

herbicides proposed to be used in the ALB program have very low toxicity and any additive 

toxicity is anticipated to be low. 

 
E–16 Appendix E. Herbicide Environmental Risk Assessment

3.0. DOSE-RESPONSE ASSESSMENT

3.1 Human Health Dose-Response Assessment

 

A dose-response assessment evaluates the dose levels (toxicity criteria) for potential human 

health effects including acute and chronic toxicity. The toxicity criteria sources include 

documents and on-line sources from the USEPA/OPP, USEPA Integrated Risk Information 

System, and the Agency for Toxic Substances and Disease Registry. If a criterion was not 

available from these sources, information in other regulatory documents or the primary 

literature was used. When toxicity criteria were developed, uncertainty factors (UFs) were 

incorporated to address data gaps, effects to sensitive groups, and variability in the study and/or 

human populations. 

As discussed in Section 2.4, triclopyr, imazapyr, and metsulfuron-methyl are not classified as 

human carcinogens. The noncancer toxicity criterion is developed by identifying a NOAEL, or 

LOAEL if an appropriate NOAEL is not available, and applying one or more uncertainty factors. 

These values are used to calculate a RfD which is a dose that will not result in any adverse 

effects to an individual. 

3.1.1 Triclopyr

The acute and chronic RfDs for triclopyr (1 and 0.05 mg/kg bw/day, respectively) are derived 

from NOAEL values from the rat studies with an uncertainty factor of 100 added for inter- and 

intraspecies differences (USEPA, 2002a; USDA FS, 2011a). The acute RfD is based on a 

developmental study with triclopyr BEE in which no effects were noted at 100 mg/kg bw/day 

but maternal toxicity was noted at 300 mg/kg bw/day. The chronic RfD is based on a twogeneration reproduction study in rats with triclopyr acid where no adverse effects were noted at 

5 mg/kg bw/day but effects on the kidney (degeneration of renal proximal tubules) were noted at 

25 mg/kg bw/day. Because of concerns for the reproductive and developmental toxicity of 

triclopyr, the chronic RfD is used to assess the risks to women of childbearing age associated 

with both acute and long-term exposure. USEPA (2002a) indicates that the acute RfD is not 

applicable to females between the ages of 13-50 years—i.e., of child bearing age. The basis for 

this recommendation appears to be signs of maternal toxicity observed at 30 mg/kg bw/day with 

a reported NOAEL of 5 mg/kg bw/day. As discussed below, the chronic RfD for triclopyr is 

0.05 mg/kg bw/day, based on a NOAEL of 5 mg/kg bw/day. Thus, for women of childbearing 

age, the EPA recommends an acute RfD of 0.05 mg/kg/day, which is equivalent to the chronic 

RfD. 

3.1.2 TCP

TCP is a metabolite of triclopyr. The acute and chronic RfDs for TCP (0.025 and 0.012 mg/kg 

bw/day, respectively) are based on NOAEL values from the rabbit and dog studies (USEPA, 

2002a, USDA FS, 2011a). The acute RfD is based on a developmental study in rabbits in which 

birth defects (an increased incidence of hydrocephaly and dilated 
Appendix E. Herbicide Environmental Risk Assessment E–17 

ventricles) chemistry were observed at the 48 mg/kg/day (LOAEL) and no effects were observed 

at the 12 mg/kg/day dose (NOAEL). 

For both acute and chronic exposures the uncertainty factor for TCP is set at 1000: 10 to account 

for uncertainties in species-to-species extrapolation and another factor of 10 to encompass 

sensitive individuals in the population, as well as an additional factor of 10 for the potential for 

increased sensitivity in children. 

3.1.3 Imazapyr

The USEPA developed a chronic RfD of 2.5 mg/kg/day for imazapyr based on a 2-year dog 

study and a reported NOEL of 25 mg/kg/day that was adjusted using an uncertainty factor of 

100 for inter- and intraspecies differences (USEPA, 2006). USEPA did not develop an 

acute/single dose RfD for imazapyr. A developmental study in rats with imazapyr reported a 

maternal NOAEL of 300 mg/kg bw/day with a LOAEL based on salivation at 1,000 mg/kg 

bw/day (see Development and Reproductive Effects under Section 2.4). Based on the maternal 

NOAEL of 300 mg/kg bw/day applying by an uncertainty factor of 100, an acute RfD will be 3 

mg/kg/day, which is similar to the chronic RfD for imazapyr . 

3.1.1 Metsulfuron-Methyl

The USEPA developed a chronic RfD of 0.25 mg/kg/day for metsulfuron-methyl based on a 2- 

year rat study and a reported NOEL of 25 mg/kg/day adjusted using an uncertainty factor of 100 

for inter- and intraspecies differences (USEPA, 2002b). USEPA did not develop an acute/single 

dose RfD for metsulfuron-methyl. A reproduction study in rats reported a parental systemic 

NOAEL of 34 mg/kg/day, with a LOAEL of 342 mg/kg/day based on decreased body weight 

(see Development and Reproductive Effects under Section 2.4). Based on the NOAEL of 34 

mg/kg bw/day and an uncertainty factor of 100, an acute RfD will be 0.34 mg/kg/day, which is 

similar to the chronic RfD for metsulfuron-methyl. 

3.1.2 Drinking Water

Triclopyr, imazapyr, and metsulfuron-methyl are not currently regulated under the Safe 

Drinking Water Act. Maximum Contaminant Levels protective of human health are not 

established for these compunds. 

3.2 Ecological Effects Analysis

This section of the risk assessment discusses available ecological effects data for terrestrial and 

aquatic biota. Available acute and chronic toxicity data are summarized for all major taxa and 

will be integrated with the exposure analysis section to characterize the risk of triclopyr, 

imazapyr, and metsulfuron-methyl to nontarget wildlife and domestic animals. Information in 

this section was gathered from on-line databases and searches for relevant peer reviewed and 

non-peer reviewed literature. Recent reviews of triclopyr, imazapyr, and metsulfuron-methyl 

by the Forest Service were used to summarize available ecotoxicological effects data. 
E–18 Appendix E. Herbicide Environmental Risk Assessment

3.2.1 Aquatic Effects Analysis 

3.2.1.1 Fish and Amphibians

Triclopyr and TCP 

The acute fish toxicity studies show that triclopyr and TCP acute toxicities to fish range from 

practically non-toxic to highly toxic (USEPA Ecotoxicity Categories, USEPA, 2014c) (table 3-

1). The acute median effective concentrations (96-hour EC50) for triclopyr TEA, triclopyr acid, 

TCP, and triclopyr BEE from the fish toxicity studies are 131 mg a.e./L, 15.3 mg a.e./L, 3.19 mg 

a.e./L, and 0.539 mg a.e./L, respectively (USDA FS, 2011a). Based on the acute toxicity data, 

triclopyr TEA is much less toxic to fish than triclopyr acid, triclopyr BEE, or TCP. Compared 

to triclopyr BEE, triclopyr TEA is less acutely toxic by a factor of approximately 240, and TCP 

is less acutely toxic by a factor of approximately 6. 

The sublethal effects of Garlon® 4 were investigated on rainbow trout using flow-through 

systems (Johansen and Green, 1990). Fish were lethargic at concentrations of 0.32-0.43 mg/L. 

Another study reported behavioral changes in rainbow trout for Garlon® 4 at 0.6 mg/L and 

Garlon® 3A at 200 mg/L (Morgan et al., 1991). 

Chronic fish toxicity studies were conducted using triclopyr TEA, triclopyr BEE, and TCP and 

are summarized in USDA FS (2011a). The 28-day toxicity study for triclopyr TEA in fathead 

minnow reported a NOEC of 32.4 mg a.e./L and a LOEC of 50.2 mg a.e./L based on effects to 

length. The chronic toxicity study for triclopyr BEE in rainbow trout reports a NOEC of 0.019 

mg a.e./L and a LOEC of 0.034 mg a.e./L based on effects to larval weight and length (USEPA, 

2009). The chronic toxicity study for TCP in rainbow trout reports a NOEC of 0.178 mg/L and 

a LOEC of 0.278 mg/L based on effects to length and weight (USDA FS, 2011a). The chronic 

fish effects data show that triclopyr BEE is the most toxic form of triclopyr and that TCP is 

more toxic than triclopyr TEA. 

Acute toxicity data for aquatic phase amphibians is limited. An acute toxicity study using the 

African clawed frog, Xenopus laevis, exposed to Garlon® 3A shows that triclopyr TEA is 

slightly toxic with a 96-hour LC50 of 84 mg a.e./L (USDA FS, 2011a). 

This LC50 is well within the range of LC50 values (≈40 to 420 mg a.e./L) for triclopyr TEA in 

fish. The acute toxicity studies in embryos and tadpoles exposed to triclopyr BEE formulations 

show triclopyr BEE is slightly toxic to amphibian embryos with a 

median 96-hour LC50 of 17.78 mg a.e./L, and moderately toxic to tadpoles with a 

median 96-hour LC50 of 2.34 mg a.e./L. The acute toxicity of triclopyr BEE to tadpoles varies 

from species to species (Rana pipiens is the most sensitive species and R. clamitans is the least 

sensitive species). Tadpoles are more sensitive than embryos by approximately an order of 

magnitude. This difference in sensitivity may reflect the rapid uptake of triclopyr BEE through 

the gills of tadpoles, relative to passive uptake by amphibian embryos (USDA FS, 2011a). 
Appendix E. Herbicide Environmental Risk Assessment E–19 

Table 3-1. Acute fish toxicity data for triclopyr acid, TEA, BEE, and TCP. 

Test Organism Endpoint/Length Toxicity Value 

(mg/L)

NOEC

(mg/L)

Triclopyr Acid

Bluegill sunfish 96-hour LC 50 155.4 NR 

Rainbow trout 96-hour LC 50 79.2 and 7.5 NR 

Chinook salmon 96-hour LC 50 9.7 NR 

Coho salmon 96-hour LC50 9.6 NR 

Chum salmon 96-hour LC 50 7.5 NR 

Sockeye salmon 96-hour LC 50 7.5 NR 

Pink salmon 96-hour LC 50 6.3 NR 

Median 15.3

Triclopyr TEA

Tidewater silverside 96-hour LC 50 40.1 NR 

Bluegill sunfish 96-hour LC 50 233.1 and 65.1 NR 

Catfish, juv 96-hour LC 50 78.3 NR 

Fathead minnow 96-hour LC 50 85.8, 86.4, 168.5, and 422.8 NR 

Chum salmon 96-hour LC50 96.1 NR 

Chinook salmon 96-hour LC 50 99 NR 

Sockeye salmon 96-hour LC 50 112 NR 

Coho salmon 96-hour LC50 167 NR 

Coho salmon, juv 96-hour LC50 127.2 NR 

Catfish, adult 96-hour LC 50 141 103 

Rainbow trout 96-hour LC 50 151, 273.7 and 286 NR 

Median 130.7

Triclopyr BEE

Bluegill sunfish 96-hour LC50 0.25 0.091 

Bluegill sunfish 96-hour LC 50 0.54 and 0.58 NR 

Coho salmon 96-hour LC 50 0.26, 0.47, and 1.0 NR 

Fathead minnow 96-hour LC50 0.5 NR 

Fathead minnow 96-hour LC 50 1.5 0.97 

Median 0.54

TCP

Rainbow trout 96-hour LC 50 1.5 and 12.6 NR 

Coho salmon 96-hour LC 50 1.8 NR 

Chum salmon 96-hour LC50 1.8 NR 

Chinook salmon 96-hour LC 50 2.1 NR 

Sockeye salmon 96-hour LC50 2.5 NR 

Pink salmon 96-hour LC50 2.7 NR 

Bluegill sunfish 96-hour LC50 12.5 NR 

Median 3.19 

NR = not reported Source: USDA FS, 2011a, Table 33 

 
E–20 Appendix E. Herbicide Environmental Risk Assessment

The acute toxicity data indicate that amphibians are less sensitive to triclopyr BEE than fish. The 

acute toxicity data in amphibians for triclopyr BEE all involve triclopyr formulations. No data 

appears to be available regarding the toxicity of unformulated triclopyr BEE or TCP to 

amphibians. The acute amphibian toxicity data for triclopyr TEA and BEE formulations are 

summarized in table 3-2. 

In addition to standardized acute toxicity studies there have been frog embryo teratogenesis 

studies conducted using Garlon® 3A and Garlon® 4 (Perkins et al., 2000). In the assay, X. laevis 

embryos were exposed to the test solution in Petri dishes for 96 hours and observed for 

malformations. No hind limb abnormalities were reported. The abnormalities observed in the 

study include uncoiling of the gut, edema, blistering, abnormal pigmentation, and axial twisting 

in control embryos. However, there were no statistically significant increases in abnormalities 

in any groups exposed to Garlon® 3A or Garlon® 4 at sublethal levels. The study results 

indicate that triclopyr at sublethal concentrations is not likely to cause reproductive or 

teratogenic effects in amphibians. 

Table 3-2. Acute amphibian toxicity data for triclopyr TEA and BEE formulations 

Test Organism Endpoint/Length Toxicity Value 

(mg a.e./L)

NOEC

(mg a.e./L)

Triclopyr TEA

African clawed frog, Xenopus laevis 96-hour LC50 84 NR 

Triclopyr BEE Formulations (Embryos) 

Xenopus laevis 96-hour LC50 13.7 NR 

Xenopus laevis 96-hour LC50 15.0 NR 

Bufo americanus 96-hour LC50 15.1 NR 

Rana pipiens 96-hour LC50 23.3 NR 

Rana clamitans 96-hour LC50 24.6 NR 

Median 17.78

Triclopyr BEE Formulations (Tadpoles) 

Rana pipiens 96-hour LC50 0.79 NR 

Bufo americanus 96-hour LC50 0.88 NR 

Xenopus laevis 96-hour LC50 1.70 NR 

Rana pipiens 96-hour LC50 2.79 NR 

Rana clamitans 96-hour LC50 3.01 NR 

Rana pipiens 96-hour LC50 3.39 NR 

Rana clamitans 96-hour LC50 11.50 NR 

Median 2.34

NR = Not reported Source: USDA FS, 2011a 

 Imazapyr 

The acute toxicity of imazapyr is classified as practically non-toxic to fish based on LC50 values 

of >100 mg a.e./L for imazapyr acid (USDA FS, 2011b). The acute toxicity for the 

isopropylamine salt of imazapyr is also practically non-toxic to fish based on acute bioassays in 

bluegill and trout. The acute toxicity for Arsenal® herbicide (27.8% a.i isopropylamine salt of 

imazapyr and 72.2% inert) is slightly toxic with a 96-hour LC50 of 41 mg a.e./L in bluegill and 

21 mg a.e./L in trout. The acute fish toxicity data for imazapyr are summarized in table 3-3. 
Appendix E. Herbicide Environmental Risk Assessment E–21 

A micronucleus assay using Tilapia rendalli (an herbivorous fish native to Africa) indicated 

positive mutagenic activity (USDA FS, 2011b; Grisolia, 2002). In this screening test, fish 

were exposed to imazapyr at 20, 40, or 80 mg/kg through intra- abdominal injections. A 

statistically significant increase in erythrocyte micronuclei was observed in the 80 mg/kg dose 

groups. However, the finding of this study does not have a substantial impact on the hazard 

identification for fish because the exposure route was atypical, and a positive response was 

only seen at the maximum dose of 80 mg/kg. In addition, imazapyr does not appear to be 

mutagenic or carcinogenic in mammals as discussed in the human health section of this risk 

assessment. 

 

Table 3-3. Acute and chronic fish toxicity data for imazapyr. 

Test Organism Endpoint/Length Toxicity Value 

(mg a.e./L)

NOEC

(mg a.e./L)

Imazapyr Acid

Bluegill sunfish 96-hour LC50 >100 NR 

Atlantic silversides 96-hour LC50 NR >184 

Bluegill sunfish 96-hour LC50 >100 NR 

Rainbow trout 96-hour LC50 >100 NR 

Channel catfish 96-hour LC50 >184 NR 

Fathead minnow Early life-stage (egg-to-fry) NR 120 

Fathead minnow Full Life Cycle NR 118 

Rainbow trout Early life-stage (egg-to-fry) NR 43.1 

(LOEC: 92.4)

Imazapyr Isopropylamine (IPA) Salt

Bluegill sunfish 96-hour LC50 >815.5 NR 

Rainbow trout 96-hour LC50 >110 110

Arsenal® Herbicide

Bluegill sunfish 96-hour LC50 40.68 NR 

Rainbow trout 96-hour LC50 20.8 10.4

NR = Not reported Source: USDA FS, 2011b 

 The USEPA/OPP risk assessment for the California red-legged frog summarized two fish kill 

incidents associated with imazapyr (USEPA, 2007). The first incident reported a 63 fish and 

algae kill in a pond 60 feet away from a mixed herbicidal spray. A mixture of the 

isopropylamine salt of imazapyr, diuron and metsulfuronmethyl was sprayed onto a fence row and may have entered the pond from drift and/or runoff. 

However, it cannot be definitively determined that the fish kill was due to exposure to imazapyr. 

The second incident involved a goldfish kill. The cause of the kill could not be determined, but 

there was suspected runoff and drift into the pond after an aerial application of an imazapyr 

formulation to a nearby 145 acres. 

The long-term toxicity of imazapyr acid to fathead minnows has been tested in an early lifestage (egg to fry) study and a full life cycle study (USDA FS, 2011b). Neither 
E–22 Appendix E. Herbicide Environmental Risk Assessment

study detected adverse effects at concentrations of up to about 120 mg a.e./L. The early lifestage chronic toxicity study in rainbow trout for imazapyr acid reported a NOEC of 43.1 mg 

a.e./L and a LOEC of 92.4 mg a.e./L due to reduced hatch and fry survival. Both acute and 

chronic toxicity studies in fish indicate that trout appear to be the most sensitive species. 

No long-term toxicity studies on imazapyr formulations have been conducted. The acute 

NOAEC of 110 mg a.e./L for the isopropylamine salt of imazapyr in rainbow trout is above the 

longer-term NOAEC of 43.1 mg a.e./L for the imazapyr acid. However, the acute NOAEC of 

10.4 mg a.e./L for the Arsenal® herbicide formulation in rainbow trout is below the longer-term 

NOAEC for the imazapyr acid. 

No information is available regarding the toxicity of imazapyr to aquatic-phase amphibians. 

Following a standard USEPA approach, fish toxicity data is used for aquatic phase amphibians 

assuming that fish are approximately as sensitive as aquatic phase amphibians. 

Metsulfuron-methyl 

Acute toxicity studies using rainbow trout and bluegill (USDA FS, 2005) show that 

metsulfuron-methyl is practically nontoxic to fish with 96-hour LC50 values >150 mg/L. 

Sublethal effects including erratic swimming behavior, laying on the bottom, lethargy and color 

changes that were observed in rainbow trout with a reported NOEC of 10 mg/L. 

The 90-day chronic exposure study in rainbow trout including fish, egg, and fry (USDA FS, 

2005) reported no effects on rainbow trout hatching, larval survival, or larval growth at a 

concentration of up to 4.7 mg/L. The study reported a LOEC of 8 mg/L with decreases in 

hatching and survival of fry. The acute and chronic fish toxicity data for metsulfuron-methyl is 

summarized in table 3-4. No information is available regarding the toxicity of metsulfuronmethyl to amphibian species. 

 Table 3-4. Acute and chronic fish toxicity data for metsulfuron-methyl. 

Test Organism Endpoint/Length Toxicity Value 

(mg/L)

NOEC

(mg/L)

Acute

Rainbow trout 96-hour LC50 >150 100 

Bluegill sunfish 96-hour LC50 >150 150 

Bluegill sunfish 96-hour LC50 >1000 1000 

Rainbow trout 96-hour LC50 >1000 10 

Chronic

Rainbow trout 90-days NR 4.7 (LOEC: 8.0)

NR = Not reported Source: USDA FS, 2005 

 
Appendix E. Herbicide Environmental Risk Assessment E–23 

3.2.1.1 Aquatic Invertebrates

Triclopyr and TCP 

Triclopyr acid and TEA are less toxic to aquatic invertebrates compared to triclopyr BEE 

(USDA FS, 2011a). TCP appears to be less toxic than triclopyr BEE, but more toxic than 

triclopyr acid and TEA to aquatic invertebrates. The aquatic invertebrate studies show that the 

triclopyr acid and TEA are practically nontoxic to non-bivalve aquatic invertebrate species with 

acute LC50 values >100 mg/L. Triclopyr TEA is slightly toxic to bivalve test species with an 

acute median LC50 of 19.7 mg/L. One study using TCP shows that TCP is slightly toxic to 

Daphnia magna with an acute LC50 of 10.9 mg/L. Triclopyr BEE and Garlon® 4 are moderately 

toxic to aquatic invertebrates with an acute median LC50 value of 2.9 mg/L. Triclopyr BEE and 

Garlon® 4 are highly toxic to bivalves with acute EC50 values between 0.1 and 1 mg/L based on 

shell deposition. The acute triclopyr and TCP toxicity data for aquatic invertebrates are 

summarized in table 3-5. 

Chronic toxicity to aquatic invertebrates has been assessed for triclopyr TEA and TCP. The 21-

day D. magna study for triclopyr TEA reported a NOEC of 25 mg a.e./L and a LOEC of 46.2 

mg a.e./L based on total number of young and mean brood size (USEPA, 2009). The chronic 

study using D. magna and TCP reported a NOEC of 0.058 mg TCP/L and a LOEC of 0.13 mg 

TCP/L based on a significant decrease in the mean number of young (USDA FS, 2011a). 

A series of 1-hour field exposure studies and triclopyr BEE in several species of stream 

invertebrates show that the LC50 values for these aquatic invertebrates were greater than 290 

mg/L (≈200 mg a.e./L) (Kreutzweiser et al., 1992). The LC50 values from the 1-hour field 

studies are two orders of magnitude higher than the standard 48-hour LC50 values for triclolpyr 

BEE suggesting those species are less sensitive. 

Imazapyr 

The acute toxicity data for aquatic invertebrates show that imazapyr acid and isopropylamine 

salt of imazapyr are practically non-toxic to D. magna (USEPA, 2005b; USEPA, 2007) and 

saltwater invertebrates—i.e., oysters and pink shrimp (USEPA, 2005b). The bioassays using D. 

magna indicate that the Arsenal® 

herbicide formulation is slighty toxic to aquatic invertebrates (more toxic than either imazapyr 

acid or the isopropylamine salt of imazapyr). The EC50 of 79 mg a.e./L for Arsenal® to D. 

magna is less than the EC50 of >100 mg a.e./L for imazapyr acid to D. magna, and the EC50 of 

614 mg a.e./L for isopropylamine salt of imazapyr (USDA FS, 2011b). The acute aquatic 

invertebrate toxicity data for imazapyr is summarized in table 3-6. 
E–24 Appendix E. Herbicide Environmental Risk Assessment

Table 3-5. Acute aquatic invertebrate toxicity data for triclopyr and TCP 

Test Organism Endpoint/Length Toxicity Value 

(mg a.e./L)

NOEC

(mg a.e./L)

Triclopyr Acid and TEA (Non-bivalve)

Grass shrimp 48-hour LC50 103.7 NR 

Daphnia magna (Acid) 48-hour LC50 132.9 NR 

Pink shrimp 48-hour LC50 270.5 NR 

Physella gyrina (Acid) 48-hour LC50 293 NR 

Daphnia magna 48-hour LC50 346, 357, and 837 NR 

Daphnia magna 48-hour LC50 376 <108 

Red swamp crayfish 48-hour LC50 6397.5 NR 

Median: 401.6

Triclopyr TEA (Bivalves)

Eastern oyster 96-hour EC50 (shell 18.4 NR 

Eastern oyster 48-hour EC50

(abnormal 

21.1 NR 

Median: 19.7

TCP 

Daphnia magna 48-hour LC50 10.9 NR

Triclopyr BEE and Garlon ®4 (Arthropods) 

Daphnia magna 48-hour LC50 0.25 and 1.2 NR 

Daphnia pulex 48-hour LC50 0.54 NR 

Grass shrimp 96-hour LC50 0.77 NR 

Grass shrimp 96-hour EC50 1.8 NR 

Red swamp crayfish 48-hour LC50 3.1 1.2 

Stonefly (Calineuria californica) 48-hour LC50 3.6 NR 

Mayfly (Ameletus sp.) 48-hour LC50 3.8 NR

Caddisfly (Brachycentrus 

americanus)

48-hour LC50 5 NR 

Daphnia magna 48-hour LC50 8.3 NR

Mayfly (Cinygma sp.) 48-hour LC50 8.95 NR 

Caddisfly (Psychoglypha sp.) 48-hour LC50 12.5 NR 

Caddisfly (Lepidostoma unicolor) 48-hour LC50 20 NR 

Median: 2.9

Triclopyr BEE and Garlon 4 (Bivalves)

Eastern oyster* (shell dep.) 96-hour EC50 0.14 0.05 

Eastern oyster (shell dep.) 96-hour EC50 0.33 NR 

Median: 0.21

*Garlon® 4 NR = not reported Source: USDA FS, 2011a 

 

 
Appendix E. Herbicide Environmental Risk Assessment E–25 

Table 3-6. Acute and chronic aquatic invertebrates toxicity data for imazapyr 

Test Organism Endpoint/Length Toxicity 

Value

(mg a.e./L)

NOEC

(mg a.e./L)

Imazapyr Acid

Daphnia magna 24-hour EC50 >100 NR 

48-hour EC50 >100 NR 

Eastern oyster 96-hour EC50 >100 132 

Eastern oyster 96-hour EC50 NR 109 (LOEC: 173) 

Pink shrimp 96-hour LC50 >189 189 

Daphnia magna 21-day LC50 >97.1 97.1 

Imazapyr IPA Salt

Daphnia magna 48-hour EC50 614 NR

Arsenal® Herbicide 

Daphnia magna 48-hour LC50 79.1 40.68 (LOEC: 81.36)

NR = Not reported Source: USDA FS, 2011b. 

 A long-term toxicity study using D. magna and imazapyr reported no effects at concentrations up 

to 97.1 mg a.e./L. Similar to fish, the chronic NOEC in daphnids is above the acute NOEC of 

40.68 mg a.e./L for Arsenal® herbicide. A mesocosm study was conducted to assess the longterm impacts of formulated imazapyr to aquatic invertebrates (Fowlkes et al. 2003). In this study, 

mixed macroinvertebrate species were exposed to mesocosms treated with Arsenal Applicators 

Concentrate® at concentrations of 0.184, 1.84, or 18.4 mg a.e./L. After a 2-week exposure period 

(comparable to the exposure period in the chronic daphnid studies), no impacts were noted on 

species richness or abundance. The NOEC of 18.4 mg a.e./L from the mesocosm study is 

consistent with the acute NOEC of 40.68 mg a.e./L for Arsenal® herbicide and the chronic 

NOAEC of 97.1 mg a.e./L in daphnids. However, USEPA considered the results of the 

mesocosm study are of limited value because potential effects at the species level were not 

examined (USEPA, 2007; USDA FS, 2011b). 

Metsulfuron-methyl 

Acute toxicity studies using D. magna show that metsulfuron-methyl is practically non- toxic 

to aquatic invertebrates with 48-hour EC50 values for immobility ranging from >150 mg/L to 

720 mg/L (USDA FS, 2005). The acute NOECs ranged from 150 to 420 mg/L. The 21-day 

chronic toxicity studies using D. magna reported NOECs ranging from 100 to 150 mg/L for 

survival, reproduction, and immobility, and a lower NOEC of 17 mg/L for growth. The acute 

and chronic aquatic invertebrate toxicity data for metsulfuron-methyl are summarized in table 

3-7. 

 
E–26 Appendix E. Herbicide Environmental Risk Assessment

Table 3-7. Acute and chronic aquatic invertebrate toxicity data for metsulfuron-methyl 

Test Organism Endpoint/Length Toxicity Value 

(mg/L)

NOEC

(mg/L)

Acute

Daphnia magna 48-hour EC50 >150 150 

Daphnia magna 48-hour EC50 720 420 

Chronic

Daphnia magna 21-days NR >150 (reproduction) 

(LOEC: >150) 

17 (growth) 

Daphnia magna 21-days NR 100 

NR = Not reported Source: USDA FS, 2005 

 

3.2.1.2 Aquatic Plants

Triclopyr and TCP 

Aquatic plant studies using blue green algae, freshwater diatoms, and green algae testing 

technical and formulated material have demonstrated that triclopyr TEA and BEE, and TCP, are 

moderately toxic (EC50 >1 to 10 mg/L) to slightly toxic (EC50 >10 to 100 mg/L) except for three 

test species (USDA FS, 2011a; USEPA, 2009). Triclopyr TEA is highly toxic to green algae 

Ankistrodesmus spp. and triclopyr BEE is very highly toxic to diatoms Navicula pelliculosa and 

Skeletonema costatum. These studies also show that triclopyr acid is the least toxic, and 

triclopyr BEE is the more toxic form of triclopyr to algae and diatoms compared to triclopyr 

TEA (table 3-8). 
Appendix E. Herbicide Environmental Risk Assessment E–27 

Table 3-8. Acute aquatic plant (algae) toxicity data for triclopyr and TCP. 

Test Organism Endpoint/Length Toxicity Value 

(mg a.e./L)

NOEC

(mg a.e./L)

Triclopyr Acid

Chlorella vulgaris 11 NR 

Chlorella pyrenoidosa 80 NR 

Kirchneria subcapitata 120-hour EC50 32.8 7 

Selanastrum capricornutum 120-hour EC50 50 22 

Triclopyr TEA

Chlorella vulgaris 8 NR 

Chlorella pyrenoidosa 54 NR 

Ankistrodesmus spp. 96-hour EC50 0.49 0.23 

Kirchneria subcapitata 96-hour EC50 5.4 8.1 

Skeletonema costatum 96-hour EC50 4.6 0.39 

Navicula pelliculosa 96-hour EC50 10.6 5.54 

Kirchneria subcapitata 120-hour EC50 12.1 NR 

Anabaena flos-aquae 168-hour EC50 4.1 1.39 

Triclopyr BEE

Anabaena flos-aquae 96-hour EC50 1.42 0.37 

Navicula pelliculosa 96-hour EC50 0.073 0.0014 

Skeletonema costatum 96-hour EC50 0.84 0.15 

Skeletonema costatum 120-hour EC50 5.9 1.0 

Kirchneria subcapitata 120-hour EC50 2.5 NR 

TCP

Kirchneria subcapitata 120-hour EC50 1.8 0.65 

Anabaena flos-aquae 120-hour EC50 1.8 0.36 

NR = Not reported. Source: USDA FS, 2011a 

 Aquatic plant studies using macrophytes have demonstrated that triclopyr acid, triclopyr TEA 

(to monocots), and triclopyr BEE (to both monocots and dicots) are moderately toxic with 7- or 

14-day median EC50 values between 1 and 10 mg/L. Triclopyr acid and triclopyr TEA are very 

highly toxic to dicots with median EC50 values <0.1 mg/L (see table 3-9). 

 
E–28 Appendix E. Herbicide Environmental Risk Assessment

Table 3-9. Aquatic plant (macrophytes) toxicity data for triclopyr 

Test Organism Endpoint/Length Toxicity Value 

(mg a.e./L)

NOEC

(mg a.e./L)

Triclopyr Acid and Triclopyr TEA

Monocots 

Lemna gibba 14-day EC50 6.06 2.5 

Lemna gibba 14-day EC50 7.6 2.5 

Lemna gibba 14-day EC50 7.8 2.5 

Lemna gibba 14-day EC50 13.58 NR 

Lemna minor 14-day EC50 15.8 NR 

Median: 9.47

Dicots 

Watermilfoil (acid) 14-day EC50 0.04 NR 

Watermilfoil (TEA) 5-week EC50 0.04 0.01 

Milfoil Hybrid 5-week EC50 0.08 0.03 

Watermilfoil (acid) 14-day EC50 0.56 NR 

Median: 0.09

Triclopyr BEE

Monocots 

Lemna gibba 14-day EC50 0.86 <0.111 

Lemna gibba 14-day EC50 6.25 NR 

Median: 2.32

Dicots 

Watermilfoil 14-day EC50 1.49 NR 

Watermilfoil 14-day EC50 4.62 NR 

Median: 2.62

NR = Not reported Source: USDA FS, 2011a 

 Imazapyr 

Imazapyr is more toxic to aquatic macrophytes than algae based on available EC50 values. 

Toxicity data for blue-green algae, green algae, and diatoms show that imazapyr is slightly toxic 

to algae (7-day EC50 ranging from 11.5 to 92 mg a.e./L) (USDA FS, 2011b). The toxicity data 

for duckweed and watermilfoil show that imazapyr is veryhighly toxic to macrophytes (7- or 14-

days EC50 ranging from 0.018 to 0.029 mg a.e./L) (USDA FS, 2011b) (table 3-10). 
Appendix E. Herbicide Environmental Risk Assessment E–29 

Table 3-10. Aquatic plants (algae and macrophytes) toxicity data for imazapyr 

Test Organism Endpoint/Length Toxicity Value 

(mg a.e./L)

NOEC

(mg a.e./L)

Algae

Imazapyr Acid

Green algae (Selenastrum 

capricornutum) 

7-day EC50 71 50.9 

Blue-green algae (Anabaena 

flosaquae) 

7-day EC50 12.2 9.6 

Freshwater diatom (Navicula 

pelliculosa) 

7-day EC50 >41 41 

Marine diatom (Skeletonema 

costatum) 

7-day EC50 92 15.6 

Imazapyr IPA Salt

Green algae (Selenastrum 

capricornutum) 

7-day EC50 11.5 7.16 

Macrophytes

Imazapyr Acid

Duckweed 

(Lemna gibba) 

14-day EC50 0.024 0.01 

Imazapyr IPA Salt

Duckweed (Lemna gibba) 7-day EC50 0.018 0.011 

Arsenal®

Water milfoil (Myriophyllum 

sibiricum)

14-day EC50 0.029 NR 

NR = Not reported Source: USDA FS, 2011b 

Metsulfuron-methyl 

Two 14-day toxicity studies using duckweed and northern watermilfoil demonstrate high 

toxicity from exposure to metsulfuron-methyl. The duckweed study reported a 14-day EC50 

value of 0.36 μg/L based on chlorosis of fronds and a NOEC value of 0.16 μg/L. The northern 

watermilfoil study reported a 14-day EC50 value of 0.22 μg/L based on a decrease in dry root 

mass (USDA FS, 2005) (table 3-11). 

The toxicity studies in algae show that metsulfuron-methyl is less toxic to algae compared to 

aquatic macrophytes. The 120-hour or 72-hour EC50 values for algae range from >95.4 to 1,560 

μg/L with effects on cell density, growth rate, and growth inhibition. The NOEC values from 

the algae studies range from 10 to 125 μg/L. 

A toxicity study using aquatic cyanobacteria reported significant growth inhibition at a 

metsulfuron-methyl concentration of 3 μg/L (Peterson et al., 1994).
E–30 Appendix E. Herbicide Environmental Risk Assessment

Table 3-11. Aquatic plants (algae and macrophytes) toxicity data for metsulfuron-methyl. 

* Ally® herbicide. NR = Not reported Source: USDA FS, 2005 

 

3.2.2 Terrestrial Effects Analysis

3.2.2.1 Mammals

Available mammalian data discussed in the human health section of this risk assessment can be 

used as a surrogate to evaluate the potential effects to wild mammals. In general mammalian 

toxicity is low for triclopyr, imazapyr, and metsulfuron-methyl. Specific information regarding 

the acute and chronic effects of each herbicide can be found in section 2.4 of this risk 

assessment. 

3.2.2.2 Birds

Triclopyr 

Triclopyr is considered practically non-toxic to slightly toxic to birds based on studies using the 

mallard and quail (USEPA, 1998; USDA FS, 2011a). The acute oral toxicity of triclopyr varies 

from 529 to 1,698 mg a.e./kg in mallard duck, and northern bobwhite quail (table 3-12). 

Reproduction studies in mallard and bobwhite quail report a NOEC of approximately 100 ppm 

and a LOEC of 200 ppm for both species. The acute dietary study reported a NOEC of 1,000 

ppm in Japanese quail. A long-term NOEC of 50 ppm was determined in Zebra finches for 

triclopyr BEE. 

Test Organism Endpoint/Length Toxicity Value 

(μg/L)

NOEC

(μg/L)

Algae

Green algae* 

(Selenastrum 

capricornutum)

Cell inhibition/120-hour EC50 NR 10 

Blue-green algae* 

(Anabaena flosaquae)

120-hour EC50 >95.4 <95.4 

Freshwater diatom* 

(Navicula pelliculosa) 

Growth rate/120-hour EC50 >95.4 95.6 

Green algae (Selenastrum 

capricornutum)

Growth inhibition/72-hour 

EC50

1,560 NR 

Green algae (Selenastrum 

capricornutum) 

Cell density/72-hour 

EC50 Area under the 

growth curve/72-hour 

EC50 Growth Rate/72-

hour EC50

372 

359 

1307 

125 

125 

125

Green algae (Chlorella 

pyrenoidosa) 

Growth inhibition/96-hour 

EC50

620 NR 

Macrophytes

Duckweed 

(Lemna gibba) 

14-day EC50 0.36 0.16 

Watermilfoil 

(Myriophyllum sibiricum)

14-day IC50 0.22 NR 
Appendix E. Herbicide Environmental Risk Assessment E–31 

Table 3-12 Acute toxicity of various forms of triclopyr to birds. 

Test Species LD50/LC50 NOEC LOEC

Acute oral/gavage

Mallard duck 

triclopyr acid 1,698 mg/kg/day 464 mg/kg bw (mortality) NA 

triclopyr TEA 1,418 mg/kg/day NA NA 

Northern bobwhite quail 

triclopyr BEE 529 mg a.e./kg bw 210 mg a.e./kg bw (mortality)/ 

≈126 mg a.e./kg bw 

(toxicity) 

≈350 mg a.e./kg bw 

TCP (99.9%) >2,000 mg a.i./kg bw 125 mg/kg bw 250 mg/kg bw 

(reduced body weight)

Acute dietary

Japanese quail 

triclopyr acid 3,272 ppm (LC50) 1,000 ppm 

(≈550 mg a.e./kg bw) 

2,000 ppm 

(≈1100 mg a.e./kg bw) 

Northern bobwhite quail 

triclopyr acid ≈2,553 mg a.e./kg bw 

(LC50) 

≈1000 mg a.e./kg bw 

(mortality) 

NA 

triclopyr TEA 5,189 ppm a.e. (≈3000 mg 

a.e./kg bw) (LC50) 

2,150 ppm formulation (995 

mg a.e./kg bw) 

NA 

triclopyr BEE 

(96.1% a.i.) 

5,401 ppm a.i. (3,885 ppm 

a.e.) (LC50) 

961 ppm a.i. (≈691 ppm a.e.) 

(toxicity) 

1,711ppm a.i. (reduced 

weight gain) 

Mallard duck 

triclopyr acid 5,620 ppm (LC50) NA NA 

triclopyr TEA 4,464.8 ppm a.e. (LC50) NA NA 

triclopyr BEE 

(93% a.i.) 

>6,689 ppm a.e. (LC50) 2,150 ppm (≈350 mg a.e./kg 

bw) 

4,640 ppm, 

reduced food 

consumption and bw gain

triclopyr BEE 

(96.1% a.i.) 

>3,885 ppm a.e. (1087.8 

mg a.e./kg bw) (LC50) 

961 ppm a.i. (246 mg a.e./kg 

bw/day) 

1,711 ppm a.i. (≈313 mg 

a.e./kg bw/day) (reduced 

body weight gain)

TCP >5,620 ppm ( LC50) NA 562 ppm (reduced body 

weight gain) 

Reproductive Studies

Mallard duck 

triclopyr acid 

(98.9% a.i.) 

100 ppm (10 mg/kg bw/day) 200 ppm (20 mg/kg 

bw/day) 

Northern bobwhite quail 

triclopyr acid 

(98.9% a.i.)

100 ppm (≈7.5 mg a.e/kg 

bw)

200 ppm (≈15 mg a.e/kg 

bw)

NA – Not available. Source: USDA FS, 2011a 

 Field studies using triclopyr applications in Forest Service programs did not cause adverse 

effects in birds (Boren et al., 1993; Schulz et al., 1992a,b). Benefits were observed in some 

cases where certain bird species benefited from changes in vegetation. 
 E–32 Appendix E. Herbicide Environmental Risk Assessment

Based on the acute gavage LD50 (>2,000 mg/kg bw) for TCP in bobwhite quail, TCP is less 

toxic than triclopyr acid, triclopyr TEA, or triclopyr BEE. The NOAEL of 125 mg/kg bw for 

TCP is similar to the NOAEL of ≅126 mg a.e./kg bw for triclopyr BEE. The dietary LC50 for 

TCP in mallards is >5,620 ppm suggesting it is practically non toxic with a reduced body 

weight gain noted at 562 ppm. 

Imazapyr 

Imazapyr has low acute toxicity to birds based on available avian studies. The acute gavage 

studies with single oral doses of imazapyr acid resulted in no signs of toxicity at a dose of 

2,510 mg a.e./kg bw in either quail or ducks (USEPA, 2007). The study using the Arsenal® 

herbicide formulation also caused no signs of toxicity at doses up to 2,150 mg formulation/kg 

bw, equivalent to about 486 mg a.e./kg bw. The acute NOAEL was the highest dose tested 

based on a lack of adverse effects. Results of these studies suggest imazapyr is considered 

practically non-toxic to birds. 

The longer-term (≈18 week) reproduction studies using the imazapyr acid indicate no 

adverse effects following exposures to dietary concentrations of up to 2,000 ppm a.e. 

(USDA FS 2011a). 

A field study reported no changes in bird populations after imazapyr was applied at about 3.7 lb 

a.e./acre for site preparation (Brooks et al., 1995). The visual surveys did not note any impacts 

on bird diversity, relative to sites treated with picloram, triclopyr, or hexazinone. Another field 

study indicated that imazapyr can improve bobwhite quail habitat by controlling hardwood 

establishment in pine stands (Welch et al., 2004). 

There was a reported bird kill incident in Aiken County, South Carolina from spraying a 

mixture of herbicides (imazapyr, diuron, and metsulfuron-methyl) on fence rows that may 

have drifted onto adjacent bird nest boxes. Imazapyr, as well as diuron and metsulfuronmethyl, were used in the incident and no definitive link to the herbicide treatments could be 

made (USDA FS, 2011b). 

Metsulfuron-methyl 

Metsulfuron-methyl has low avian acute toxicity based on acute and subchronic toxicity 

studies conducted in bobwhite quail and mallards (USDA FS, 2005). The 14-day oral LD50 of 

technical grade metsulfuron-methyl administered by gavage in adult bobwhite quail was 

>2,250 mg/kg. The 5-day dietary exposure of juvenile ducks and quail to technical grade 

metsulfuron-methyl at concentrations ranging from 292 to 5,620 ppm did not result in any. 

observable toxicity. The only sign of toxicity following a 5-day dietary metsulfuron-methyl 

exposure of 3,160 ppm to 10-day-old bobwhite quail and 5,620 ppm to 8-day old mallard was 

weight loss. The NOAELs for weight loss ranged from 1,780 to 3,160 ppm in both studies. 

Two 23-week feeding studies on reproductive effects for metsulfuron-methyl in bobwhite 

quail and mallards reported a NOAEL of 1,000 ppm (the highest dose tested) for chronic 

dietary exposure (USDA FS, 2005). 
Appendix E. Herbicide Environmental Risk Assessment E–33 

3.2.2.3 Reptiles and Amphibians (terrestrial phase) 

No information regarding the toxicity of imazapyr to reptiles or terrestrial-phase amphibians 

was identified in the open literature or in studies submitted to the USEPA (USEPA, 2005b; 

2006; 2007). The USEPA ecological risk assessment for imazapyr includes a risk evaluation 

to terrestrial phase amphibians using birds as a surrogates for terrestrial phase amphibians 

and reptiles (USEPA, 2005b). The same approach is used in the USEPA risk assessment for 

the California red legged frog (USEPA, 2007). The permeability of amphibian skin to 

pesticides and other chemicals is a concern for this approach. No data is available on the 

permeability of amphibian skin to imazapyr. 

Quaranta et al. (2009) noted that the skin of the frog R. esculenta is much more permeable to 

three pesticides (atrazine, paraquat, and glyphosate) than the skin of pig ear due to the 

differences in the structure and function of amphibian skin relative to mammalian skin. 

An open literature search did not identify any reptile toxicity studies using metsulfuronmethyl but identified one study on the toxicity of metsufluron-methyl to terrestrial-phase 

amphibians (Lajmanovich et al., 2013). The study determined the individual toxicity of 

herbicide formulations contining glyphosate, metsulfuron-methyl, bispyribac-sodium, and 

picloram, and compared the mixture toxicity of three binary combinations (glyphosate and 

metsulfuron-methyl, glyphosate and bispyribac-sodium, glyphosate and picloram) on 

Rhinella arenarum tadpoles. The 48 hour LC50 for metsulfuron-methyl is 

105.56 mg a.i./L. The NOEC and LOEC for metsulfuron-methyl were 80 mg a.i./L and 160 

mg a.i./L, respectively. 

3.2.2.4 Terrestrial Invertebrates and Microorganisms

Terrestrial Invertebrates 

USEPA/OPP (1998) classifies triclopyr as practically non-toxic to bees based on the results of 

acute contact toxicity studies in honey bees using triclopyr acid and triclopyr TEA (contact 

LD50 values >100 μg/bee). The contact LD50 for triclopyr BEE to the honey bee is reported as 

>72 μg/bee suggesting low toxicity (USEPA, 2009). 

The acute (14 day) toxicity studies in earthworms using triclopyr, triclopyr TEA, triclopyr 

BEE, and Garlon® 4 indicate that triclopyr acid is the least toxic to earthworms with an 

NOAEL of approximately 790 ppm a.e. ( i.e., 790 mg a.e./kg soil dry weight), and an LC50 of 

1,110 ppm a.e. (USDA FS, 2011a). Triclopyr TEA is the most toxic form of triclopyr to 

earthworms with an LC50 of about 146 ppm a.e., and an LOAEC of 134 ppm a.e. based on a 

significant increase in mortality (35% relative to 0% in the control groups) and a significant 

decrease in body weight (17%) relative to the control group. 

This study suggests that triclopyr TEA may be moderately toxic to earthworms relative to 

triclopyr acid. 

A chronic 56-day earthworm bioassay testing Garlon® 4 showed no adverse effects on 

reproduction or growth at a concentration of 6.9 ppm a.e. (USDA FS, 2011a). The study on 

the impacts of Garlon® 3A to earthworms and other invertebrates at an application rate 
 E–34 Appendix E. Herbicide Environmental Risk Assessment

of 0.56 kg a.i./ha (≈0.36 lb a.e./acre) to turf plots showed consistent results with no 

significant reduction in mixed earthworm populations, mites, springtails, or ants in turf and 

soil core samples (Potter et al. 1990). 

 

A series of field studies using triclopyr in broadcast applications suggested some effects to 

terrestrial invertebrates including beetles, butterflies, and spiders (USDA FS, 2011a). These 

effects are considered secondary because they were attributable to changes in vegetation cover, 

similar to changes in invertebrate populations observed with other vegetation management 

methods, instead of the potential toxic effect of triclopyr. 

 

Imazapyr has low acute toxicity to to the honey bee based on the acute contact LD50 of greater 

than 100 μg/bee, which is equivalent to about 860 mg/kg bw (USEPA, 2005b; 2007). There is 

no information available on other potential subchronic or non-lethal effects in bees or other 

invertebrate species. 

Honey bee toxicity studies using metsulfuron-methyl indicate that the acute LD50 is greater than 

25 μg/bee and possibly greater than 100 μg/bee, suggesting very low toxicity (USDA FS, 2005). 

These values correspond to doses ranging from about 270 to 1,075 mg/kg (0.025 mg/0.000093 

kg to 0.1 mg/0.000093 kg) using a body weight of 0.093 g for the honey bee (USDA APHIS, 

1993). 

A toxicity study on three insect species (large white butterfly, beetle, and grain aphid) placed on 

plants sprayed with metsufuron-methyl at an application rate of 0.00004 to 

0.003 lbs a.i./acre showed no adverse effects on survival or larval growth rate (Kjaer and 

Heimbach, 2001). 

Terrestrial Microorganisms 

Triclopyr is unlikely to have an impact on soil microorganisms (i.e., soil microbial function or 

community structure) based on studies on growth inhibition of fungi species (Chakravarty and 

Sidhu, 1987; USDA FS, 2011a). A study of Garlon® 4 (Estok et al., 1989) indicated a significant 

reduction of radial growth in fungi species at concentrations ≥1,000 ppm with total growth 

inhibition at ≥5,000 ppm. The slowest growing fungus was the least sensitive to triclopyr. 

Triclopyr TEA did not show any impacts on soil microbial function or community structure at an 

application rate of 1.9 kg a.i./ha (Houston et al., 1998). 

Information on effects on terrestrial microorganisms is not available for imazapyr (USEPA, 

2005b; 2007). Metsulfuron-methyl studies showed that at a concentration of 5 ppm in culture 

inhibited the growth of several strains of Pseudomonas (USDA FS, 2005). The inhibition of 

terrestrial microorganisms was attributed to ALS inhibition because terrestrial microorganisms 

have an enzyme involved in synthesis of branched chain amino acids. This enzyme is 

functionally equivalent to the target enzyme in terrestrial macrophytes. 

 
Appendix E. Herbicide Environmental Risk Assessment E–35 

Terrestrial Plants

Triclopyr 

Triclopyr causes uncontrolled abnormal growth in plants by mimicking indole auxin plant 

growth hormones. It is a selective herbicide that is most toxic to broadleaf plants (Lewer and 

Owen, 1990). Triclopyr is effective in the control of dicots and relatively ineffective in 

controlling monocots through foliar application (USDA FS, 2011a). 

The vegetative vigor studies in non-target plants from direct foliar application to young plants 

indicate that triclopyr TEA and triclopyr BEE have similar toxicities to a range of species 

(sunflower, sugar beet, tomato, oilseed rape, radish, soybean, wheat, corn, and onion for 

triclopyr TEA, and alfalfa, carrots, corn, oats, onions, radishes, soybeans, sunflowers, tomatoes, 

and wheat for triclopyr BEE). The sunflower (a dicot) is the most sensitive species for both 

TEA (EC25 = 0.005 lb a.e./acre) and triclopyr BEE (EC25 

≈0.0064 lb a.e./acre). Wheat and oats (monocots) are much more tolerant with EC25 values in 

excess of 0.3 lb a.e./acre for triclopyr TEA and triclopyr BEE (USDA FS, 2011a). The range of 

NOECs for triclopyr TEA ranges from 0.0028 lb a.e./acre testing sunflowers and soybeans to 

0.23 lb a.e./acre for barley based on effects to shoot length (USEPA, 2009). The NOECs in 

alfafa, carrots, corn, oats, onions, radishes, soybeans, sunflowers, tomatoes, and wheat from 

triclopyr BEE exposure, based on shoot length and shoot weight effects, range from 4.4 g a.i./ha 

(sunflowers for shoot length, and carrots and sunflowers for shoot weight) to >2,242 g a.i./ha 

(oats). The NOECs from another study using triclopyr BEE were <0.063 lb a.e./acre (onions) 

and 0.028 lb a.e./acre (sunflowers) (USEPA, 2009). 

The seedling emergence studies indicate that triclopyr BEE is about equally effective against 

dicots and at least some monocots. Triclopyr BEE is much more toxic than triclopyr TEA in 

some species. For example, the EC25 in alfafa for triclopyr BEE is 40 g a.i./ha (≈0.02 lb 

a.e./acre), which is much lower than the EC25 values for triclopyr TEA for all tested species 

(>0.23 lb a.e./acre). The higher phytotoxicity of triclopyr BEE may relate to faster absorption 

compared to triclopyr TEA that has been reported in chickweed, wheat, and barley (Lewer and 

Owen 1990). Variation in species sensitivity to triclopyr BEE may directly relate to the plant’s 

rate of metabolic ester hydrolysis (Lewer and Owen, 1990). 

Field studies using triclopyr have shown some adverse impacts at high application rates and 

using broadcast applications. The field studies in various conifer species show that these 

species tend to be tolerant to triclopyr exposure in spring (April and May) and after fall 

dormancy (September) (USDA FS, 2011a). 

Exposure to triclopyr drift may cause long-term impacts on some bryophyte and lichen 

communities (Newmaster et al., 1999). Some Forest Service events suggested that 

volatilization of triclopyr BEE may damage nontarget plants if it is applied under a poorly 

ventilated canopy and high temperatures. However, none of the field studies involving 

triclopyr BEE documented damage to nontarget plant species through volatilization (USDA 

FS, 2011a). 
 E–36 Appendix E. Herbicide Environmental Risk Assessment

Imazapyr 

Imazapyr inhibits acetolactate synthase, an enzyme that catalyzes the biosynthesis of branchedchain amino acids essential for protein synthesis and plant growth (USDA FS, 2011b). After 

foliage application, imazapyr is rapidly transported by the phloem from the treated leaves to the 

roots . Imazapyr has the potential to induce allelopathic effects; however, the potential for 

allelopathic effects may not substantially increase the risk to non-target plants because its 

movement in soil is relatively rapid (USDA FS, 2011b). 

The bioassays for vegetative vigor (i.e., post-emergence applications) and seedling emergence 

(i.e., pre-emergence applications) in response to foliar applications of imazapyr indicate that 

dicots are more sensitive to imazapyr than monocots. The differences between dicots and 

monocots are greater in the vegetative vigor assays compared to the seedling emergence studies. 

In the vegetative vigor assays, the EC25 of the most sensitive dicot, cucumber (0.0009 lb 

a.e./acre) is about a factor of 13 below the EC25 of the most sensitive monocot, wheat (0.012 lb 

a.e./acre). The NOAECs for the acid (22.6% a.e.) for dicots range from 0.000064 lb a.e./acre in 

cucumber to 0.0039 lb a.e./acre in sunflower, and for monocots range from 0.0039 lb a.e./acre in 

wheat and oat, to 0.0078 lb a.e./acre in corn. The NOAECs for the isopropylamine salt for 

dicots are 

0.001 lb a.e./acre in sugar beet and 0.008 lb a.e./acre in soybean, and for monocots is 

0.005 lb a.e./acre in onion. In seedling emergence assays, the EC25 of the most sensitive dicot, 

sugar beet (0.0024 lb a.e./acre) is a factor of approximately 2 below the most sensitive monocot 

species, wheat (0.0046 lb a.e./acre). The NOAECs for imazapyr acid (22.6% a.e.) in dicots are 

0.00017 lb a.e./acre in sugar beet and 0.0003 lb a.e./acre in tomato, and in monocots range from 

0.00099 lb a.e./acre in wheat to 0.0156 lb a.e./acre in oat. For the most sensitive species of dicot, 

foliar applications (vegetative vigor) are more toxic than soil exposures (seedling emergence) by 

about a factor of about 2.6, and vice versa for the most sensitive species of monocot (USDA FS, 

2011b; USEPA, 2005b). 

Metsulfuron-methyl 

Metsulfuron-methyl inhibits acetolactate synthase, an enzyme that catalyzes the biosynthesis 

of three branched-chain amino acids (valine, leucine, and isoleucine) essential for plant 

growth. 

Laboratory toxicity studies (pre- and post-emergence bioassays) using metsulfuron- methyl to 

evaluate the effects to terrestrial plants by a direct application has been evaluated for 10 species 

of plants including both dicots (soybean, cocklebur, cotton, morningglory, wild buckwheat, and 

sugar beet) and monocots (corn, barnyardgrass, rice, and nutsedge) (USDA FS, 2005). The 

study results showed that the most sensitive species was the morningglory with 70% growth 

inhibition at pre-emergence applications of about 0.00022 lbs a.i./acre. At the same application 

rate, the cocklebur and sugar beet showed 20% and 40% growth inhibition, and rice was the only 

monocot to respond with 20% inhibition. All of the plants showed 60 to 100% growth inhibition 

at an application of 0.014 lbs a.i./acre which is approximately a factor of two below the typical 

application rate used by the Forest Service (0.03 lb/acre). 

Other laboratory toxicity studies on pre-emergence and post-emergence were performed in corn, 

cucumber, onion, pea, rape, sugar beet, sorghum, soybean, tomato, and wheat 
Appendix E. Herbicide Environmental Risk Assessment E–37

(USDA FS, 2005). The pre-emergence assay showed that the most sensitive species were 

cucumber and onion with a NOEC of 0.000037 lb/acre, and the most tolerant species was wheat 

with a NOEC of 0.0056 lb/acre. The post-emergence assay showed that the cucumber was also 

the most sensitive species, with a NOEC of 0.000037 lb/acre, and the most tolerant species was 

wheat with a NOEC of 0.0039 lb/acre. 

Field studies have shown that the lowest application rate associated with adverse effects (a 

decreased yield of both tomatoes and onions) was 0.1 g/ha (USDA FS, 2005; Obrigawitch et al., 

1998), which is similar to the LOEC of 0.25 g/ha reported for other dicots. The most tolerant 

species in the field studies (Obrigawitch et al., 1998) consisted of various grasses for which 

NOEC values based on crop yield ranged up to 6 g/ha (NOEC values for wheatgrass and 

bromegrass). The NOEC value of 6 g/ha is equivalent to an application rate of about 0.0054 

lb/acre, which is identical to the NOEC value for wheat (0.0056 lb/acre). 

4.0 EXPOSURE ASSESSMENT

4.1 Human Health Exposure Assessment

Exposure assessments estimate the potential exposure of humans to triclopyr, imazapyr, and 

metsulfuron-methyl. The exposure assessment begins with the use and application methods of 

triclopyr, or triclopyr mixed with imazapyr and metsulfuron-methyl products. An identified 

exposure pathway for each herbicide includes (1) a release from a source, (2) an exposure point 

where contact can occur, and (3) an exposure route such as ingestion, inhalation, or dermal 

contact by which contact can occur (USEPA, 1989). Exposures for the identified human 

populations are qualitatively evaluated for each identified exposure pathway. 

4.1.1 Identification of Potentially Exposed Human Populations 

and Complete Exposure Pathways

Under the expected use pattern and the proposed application methods (manual painting or direct 

spray treatment using a backpack sprayer), workers in the program who mix and apply the 

herbicides in the field are the most likely human population with the potential to be exposed to 

these compounds. Dermal exposure may occur from both the manual painting application and 

backpack spray application as well as and spill/splashing-during mixing and loading herbicide 

into the containers used during application. Exposure during transportation is not anticipated 

because the container of the concentrated material is sealed. Following label directions including 

the use of proper PPE will minimize exposure to workers. 

Accidental exposure may occur during mixing and applying the formulations. Under an 

accidental spill scenario, workers may be exposed to triclopyr, imazapyr, and metsulfuronmethyl through dermal contact. However, the potential dermal contact exposure would be 

limited because these accidental events would be of low frequency and short duration. 

 
 E–38 Appendix E. Herbicide Environmental Risk Assessment

In addition to worker exposure, there is the possibility of exposure to residents in areas after 

treatment. The general public is not expected to be exposed to the herbicides during the 

application process. Exposure of residents whose property has been treated can be minimized 

through proper notification prior to treatment. However, there is potential for the general public 

to be exposed to herbicides shortly after application by walking through a treated area and sitting 

on treated stumps. Exposure to the general public to triclopyr residues was evaluated. Sensitive 

members of the population (i.e., adult females and pre-teen and teenaged children (age 11 to 16)) 

were used for the general public exposure assessment. Dermal contact with treated tree surfaces 

is the primary exposure route for the general public. Incidental ingestion and inhalation exposure 

routes are highly unlikely and were not evaluated. A dietary exposure route to the public is not 

expected because all of the herbicides will be used to only to treat tree stumps or sprouting 

vegetation. 

A significant exposure pathway is not idenified for groundwater or surface water in proximity of 

herbicide-treated stumps. Although triclopyr, imazapyr, and metsulfuron- methyl are all soluble, 

mobile, and can leach from soil to groundwater, the leaching potential is limited because the 

herbicides are applied directly to the tree stump or associated sprouting vegetation, which 

minimizes contact with the soil. Although surface runoff from treated stumps into aquatic 

resources is possible, the use pattern of spot treatment, current label restrictions, and low 

application rates limit the amount of these herbicides runoff to surface water. 

4.1.2 Exposure Evaluation

This section quantitatively evaluates worker exposure from direct contact while applying 

triclopyr using direct stump applications and backpack spraying, as well as to the general public 

from exposure resulting from direct contact with contaminated vegetation. The exposure 

scenarios evaluated for workers and the general public are related to short-term exposures 

because these herbicides are not persistent in the environment and probability of long-term 

exposure is low. 

Workers: The general worker exposure evaluated in this assessment is based on the two 

application methods, manual painting and backpack direct spray. For the manual painting 

application, the direct contact exposure scenario to a triclopyr solution is characterized by 

wearing triclopyr-contaminated gloves for 1 minute and 10 minutes. It is assumed that wearing 

gloves grossly contaminated with a chemical solution is equivalent to immersing the hands in the 

pesticide solution. Therefore, the chemical concentration in contact with the skin and the dermal 

absorption rate are essentially constant. Because the concentration of triclopyr in contact with the 

skin is nearly constant and the rate of absorption is constant, zero-order absorption (i.e., the 

dermal absorption rate is constant over time) kinetics is used. The rate of absorption under this 

exposure scenario is estimated based on central estimate (range) dermal permeability coefficients 

(Kp) of 1.3 x 10-4 (7.8 x 10-5 – 2.3 x 10-4) cm/hour 

for triclopyr TEA and 8.3 x 10-3 (4.4 x 10-3 – 1.6 x 10-2) cm/hour for triclopyr BEE (USDA FS, 

2011c,d). 

For the backpack spray application, the direct contact exposure scenario estimate for a backpack 

worker to a triclopyr solution are characterized based on the amount of material 

handled per day. The worker exposure rate is 0.01 mg/kg bw per lb and ranges between 
Appendix E. Herbicide Environmental Risk Assessment E–39

0.001 mg/kg bw per lb and 0.08 mg/kg bw per lb. The worker exposure rates are thegeometric 

mean and 95% confidence intervals, based on a log-transformation of data from a Forest Service 

assessment of worker exposure rates for triclopyr BEE (USDA FS, 2012). 

For the spill/splash related dermal contact exposure scenario, exposure to a triclopyr solution is 

characterized by a spill on the lower legs as well as the hands, with a certain amount of the 

chemical adhering to the skin. The absorbed dose is calculated as the product of the amount of 

chemical on the surface of the skin (i.e., the amount of liquid per unit surface area multiplied by 

the surface area of the skin over which the spill occurs and the chemical concentration in the 

liquid), the first-order dermal absorption rate coefficient (Ka) (expressed as a proportion of the 

deposited dose absorbed per unit time—e.g., hour-1), and the duration of exposure. The rate of 

absorption under this exposure scenario is estimated based on the first-order dermal absorption 

rate coefficient (Ka) of 6.7 x 10-4 (2.4 x 10-4 – 1.8 x 10-3) hour-1 for triclopyr TEA and 2.6 x 10-3 

(1.0 x 10-3 – 6.7 x 10-3) hour-1 for triclopyr BEE (USDA FS, 2011c,d). Details regarding exposure 

assumptions are presented in Attachment A. 

General Public: The exposure scenario developed for the general public assumes that an 

individual is exposed to the compound shortly after its application. The exposure scenario from 

contact with contaminated tree stumps assumes that the herbicide is sprayed at a given application 

rate, and that an adult female or a pre-teen and teenage child comes in contact with the sprayed 

tree stump surfaces on the same day. This exposure scenario depends on estimates of 

dislodgeable residue (the estimated amount of the chemical which could be released from 

vegetation, expressed in units of pesticide mass/surface area of vegetation), and dermal transfer 

rates (the rate at which the chemical is transferred from the contaminated vegetation to the surface 

of the skin). Dislodgeable residues are based on the pesticide, the formulation, and the sitespecific conditions. Dermal transfer rates are reasonably consistent for a number of pesticides 

(Durkin et al.1995). This exposure assessment assumes a contact period of 10 minutes and that 

the chemical is not effectively removed by washing within 1 or 2 hours of exposure. The 

assumptions regarding body weight, skin surface area, and first- order dermal absorption rates are 

summarized in Attachment A. 

4.2 Ecological Exposure Assessment

4.2.1 Aquatic Exposure Assessment

The three herbicides proposed for use in the ALB Eradication Program exhibit properties that 

suggest they could contaminate surface water. High solubilty and low soil adsorption coefficients 

increase the likelihood of off-site transport from runoff. Drift is not anticipated to be a significant 

pathway to aquatic environments from the proposed treatments. 

Herbicides are either hand painted onto stumps or backpack applications are made directly to the 

the stump and associated vegetation. The use of backpack sprayers, large coarse droplets during 

application, and other label requirements will result in negligible drift. Estimates of potential 

aquatic residues from runoff were evaluated for triclopyr BEE since it has a greater toxicity to 

aquatic vertebrates and invertebrates compared to triclopyr TEA, imazapyr, and metsulfuronmethyl. Aquatic residue values for this assessment were generated using the GLEAMS 

environmental fate model (USDA FS, 2003). The Forest Service determined the 

residues of triclopyr BEE in a small stream for three soil types and average annual rainfall 
 E–40 Appendix E. Herbicide Environmental Risk Assessment

amounts ranging from 5 to 250 inches per year. The application rate was1 lb a.e./acre with 

resulting residues ranging from zero to 0.149 mg/L. The highest concentrations were observed at 

the maximum average annual rainfall levels and in sandy soils which would be highly permeable 

and susceptible to runoff. The residues from these estimates were then compared to the available 

aquatic effects data for triclopyr BEE which is further discussed in the risk characterization 

section of this risk assessment. 

4.2.2 Terrestrial Exposure Assessment

The potential for herbicide exposure to non-target terrestrial vertebrate wildlife is very low under 

the proposed uses. Treatments to stumps would not result in any appreciable exposure to nontarget terrestrial vertebrates through either the oral, dermal, or inhalation pathways. In addition, 

the available toxicity data for birds and mammals through these exposure pathways demonstrates 

very low toxicity. There is the possibility that some mammals may graze on treated sprouting 

vegetation; however, there is no plausible exposure scenario resulting in adverse effects based on 

daily food consumption rates. There is the potential for exposure from drinking water. However, 

estimates for triclopyr BEE, discussed above, are very low and any non-target vertebrate wildlife 

would have to consume several times their daily water intake to receive a dose that would 

approach an adverse effect. 

 

5.0 RISK CHARACTERIZATION

5.1 Human Health

Risks associated with adverse human health effects are characterized quantitatively in this 

section. The potential adverse health effects to subgroups within the population from exposure 

to triclopyr were evaluated using a hazard quotient (HQ) approach. A HQ is the ratio of the 

estimated level of total exposure to the relevant noncancer toxicity criterion (i.e., appropriate 

RfD) for a specific exposure pathway (or for a single substance). If there is more than one 

chemical, the HQs are summed over all exposure pathways and all chemicals to develop a total 

hazard index (HI). If the HI is less than 1, it is considered unlikely that the exposure will cause 

adverse health effects; conversely, if the HI is greater than 1, then health effects may result from 

exposure. Unlike cancer risk estimates that are expressed as probabilities (e.g., 1 in a million), 

HIs are deterministic methods of estimating risk and do not represent the probability of health 

effects, but are presented to provide context to the relationship between the known toxicity of a 

substance and the estimated magnitude of exposure. The risk estimations associated with the 

triclopyr exposure scenario are expected to be representative of the mixed herbicide 

scenario. This is because triclopyr has at least an order of magnitude higher toxicity compared to 

the other two herbicides (RfD of 0.05 mg/kg/day for triclopyr, RfDs of 2.5 mg/kg/day for 

imazapyr, and 0.25 mg/kg/day for metsulfuron-methyl). Adding HQs from the other two 

compounds is not expected to raise the total HI to exceed the USEPA level of concern (HI of 

1). 

Workers: The HQs for general and accidental exposures of workers are calculated and 

summarized in table 5-1. The HQs are based on the total exposure that a worker might receive 

during a directed foliar application at a label allowable maximum application rate of 9 lb/acre 

(TEA) and 8 lb/acre (BEE) with the acute RfD of 1 mg/kg/day. The use of the acute RfD for 
Appendix E. Herbicide Environmental Risk Assessment E–41

risk characterization for the general worker exposure (male workers) is intended only to 

illustrate the consequences of applying triclopyr sporadically as part of other activities. The HI 

values for the three worker-exposure scenarios (i.e., manual painting application direct contact, 

backpack spray application direct contact, and spill/splashing-related dermal contact) are less 

than or do not exceed the USEPA level of concern of 1 under the central and lower estimations 

for both triclopyr TEA and triclopyr BEE. The upper estimate HQ of 7 for triclopyr BEE under 

the contaminated gloves with 10 minutes exposure scenario and the upper estimate HQs of 6 for 

triclopyr TEA and 5 for triclopry BEE under the backpack application exposure scenario 

exceeded the level of concern. However, these risk estimations are conservative values based on 

the maximum labelled application rates (9 lb a.e./acre TEA and 8 lb a.e./acre BEE) instead of a 

typical application rate (1 lb a.e./acre). The hazard quotients calculated for a typical application 

rate are less than one (table 5-2). The backpack application exposure scenario also 

conservatively assumes no protective clothes resulting in overestimation of risk, which will be 

minimized by the proper use of PPE. The central risk estimates are a more typical exposure 

scenario. Therefore, it is unlikely that the exposure to workers will cause adverse health effects. 

 Table 5-1. Hazard quotient values estimated for workers making triclopyr applications at 

maximum rates. 

Hazard Quotient 

Exposure Scenario Form Central 

Estimate 

Lower 

Estimate 

Upper Estimate 

Accidental/Incidental Exposure

Contaminated Gloves, 1 min. TEA 2 x 10-3 5 x 10-4 1 x 10-2

Contaminated Gloves, 10 mins 

Spill on Hands, 1 hour 

BEE 

TEA 

BEE 

TEA 

7 x 10-2

1 x 10-2

7 x 10-1

3 x 10-3

2 x 10-2

5 x 10-3

2 x 10-1

7 x10-4

7 x 10-1

1 x 10-1

7

4 x10-2

Spill on lower legs, 1 hour 

BEE 

TEA 

BEE 

1 x 10-2

7 x 10-3

2 x 10-2

3 x10-3

1 x10-3

6 x10-3

1 x10-1

9 x10-2

3 x10-1

General Exposures

Backpack Applications TEA 4 x 10-1 1 x10-3 6

BEE 4 x 10-1 1 x10-3 5

HI TEA 0.4 0.008 6

BEE 1 0.2 12

Bold and shaded - Hazard quotient/index exceeding the target hazard quotient/index of 1. 

Note: HI is the sum of HQs of the contaminated gloves (10 minutes), spill on hands, spill on lower legs, 

and backpack applications exposure scenarios. 
 E–42 Appendix E. Herbicide Environmental Risk Assessment

Table 5-2. Hazard quotient values estimated for workers making triclopyr applications at 

typical application rate. 

Hazard Quotient 

Exposure Scenario Form Central 

Estimate 

Lower 

Estimate 

Upper Estimate 

Accidental/Incidental Exposure

Contaminated Gloves, 1 min. TEA 2 x 10-3 5 x 10-5 1 x 10-3

Contaminated Gloves, 10 mins 

Spill on Hands, 1 hour 

BEE 

TEA 

BEE 

TEA 

9 x 10-3

1 x 10-3

9 x 10-2

3 x 10-4

3 x 10-3

5 x 10-4

3 x 10-2

8 x10-5

8 x 10-2

1 x 10-2

8 x 10-1

5 x10-3

Spill on lower legs, 1 hour 

BEE 

TEA 

BEE 

1 x 10-3

7 x 10-4

3 x 10-3

3 x10-4

2 x10-4

7 x10-4

2 x10-2

1 x10-2

4 x10-2

General Exposures

Backpack Applications TEA 4 x 10-2 2 x10-4 6 x 10-1

BEE 4 x 10-2 2 x10-4 6 x 10-1

 

General Public: The HQs for potential herbicide exposure to the general public are calculated 

and summarized in table 5-3. The HQs are based on the exposure scenario where an adult 

female or a pre-teen and teenaged child (age 11-16) could be exposed to dislodgeable residues 

shortly after a foliar application at the label allowable maximum application rates of 9 lb/acre 

(TEA and TCP) or 8 lb/acre (BEE), for a direct contact time of 10 minutes and exposure duration 

for 1 hour (adult female) and 2 hour (pre-teen and teenaged child). The chronic RfDs of 0.05 

mg/kg/day (TEA and BEE) and 0.012 mg/kg/day (TCP) were used for the adult female as a 

conservative approach. The HIs for the adult female and pre-teen and teenaged child are all less 

than USEPA level of concern of one suggesting that it is unlikely that herbicide exposure to the 

general public will cause adverse health effects. 

Table 5-3. Hazard quotient values estimated for general public after triclopyr applications. 

Exposure Scenario 

 Hazard Quotient 

Form Central Estimate Lower Estimate Upper Estimate 

Non-accidental Acute Exposure 

Vegetation contact, shorts and T-shirt – an adult female 

 TEA 3 x 10-3 1 x 10-3 8 x 10-3

 BEE 1 x 10-2 4 x 10-3 2 x 10-2

 TCP 2 x 10-1 7 x 10-2 6 x 10-1

HI 0.2 0.08 0.6 

Vegetation contact, shorts and T-shirt – a pre-teen and teenaged child (11-16)

 TEA 3 x 10-4 1 x10-4 8 x10-3

 BEE 1 x 10-3 4 x10-4 3 x10-3

 TCP 2 x 10-1 8 x10-2 6 x10-1

HI 0.2 0.08 0.6 
Appendix E. Herbicide Environmental Risk Assessment E–43

5.2 Aquatic and Terrestrial Ecological Risk Characterization

This section of the risk assessment will integrate the effects analysis regarding toxicity studies 

summarized for terrestrial and aquatic nontarget organisms with the potential for exposure to 

quantify whether direct or indirect risks would be anticipated for nontarget organisms and 

domestic animals from program use of triclopyr, imazapyr, and metsulfuron-methyl. Direct risk 

refers to those risks that could occur from direct exposure to the herbicides while indirect risks 

refer to impacts that could occur to prey or habitat that nontarget organisms rely on for food and 

shelter. In cases where the range of response data to each herbicide does not fall within the 

range of potential exposure values, impacts to individuals and populations are not anticipated. 

Further evaluation of the assumptions used in the risk characterization is required to refine risk 

where residues exceed the response data. 

5.2.1 Aquatic

The available aquatic toxicity data for vertebrates and invertebrates for all three herbicides 

suggests that these taxa would be at very low risk from herbicide applications. Triclopyr BEE 

poses the greatest risk to aquatic vertebrates and invertebrates based on its toxicity profile while 

imazapyr and metsulfuron-methyl pose the least amount of risk because both products would be 

considered practically non-toxic to aquatic invertebrates and vertebrates. Based on the low 

frequency of use in the ALB program and the method of application, there is a low likelihood that 

any associated drift and runoff from these applications would result in significant risk to aquatic 

vertebrates or invertebrates. The risk to aquatic plants is greater based on the known toxicity for 

each herbicide. Effects to aquatic plants can result in indirect effects to aquatic vertebrates and 

invertebrates that rely on these species for food and shelter. The proposed use pattern would 

minimize the potential for significant residues from drift or runoff to aquatic habitats. In addition, 

current label requirements restricting applications in aquatic areas would further reduce the 

potential for residues. 

The acute aquatic risk from the use of triclopyr BEE was determined for aquatic vertebrates, 

invertebrates and plants based on a typical application rate (1 lb a.e./acre) and the maximum 

labelled used rate (8 lb a.e./acre) (figure 5-1) 

Residue values were compared to the range of acute aquatic toxicity data. A presumption of risk 

occurs when there is an overlap between residues and the range of acute toxicity data. This was 

done for triclopyr BEE due to its ten fold or greater toxicity when compared to triclopyr TEA, 

imazapyr, and metsulfuron-methyl. The waterbody that was modeled in this exercise was a small 

stream (USDA FS, 2003). Residues were estimated using the GLEAMS environmental fate model 

for three soil types (clay, loam, and sand) and annual yearly rainfall amounts ranging from 5 to 250 

inches. Chronic risk comparisons were not made because triclopyr BEE degrades rapidly in water 

to triclopyr acid, and chronic exposure to the ester would not occur. The residues estimated from 

this excerise are conservative estimates because they assume applications occur to soil, which 

would not be the case for proposed stump applications. There is the possibility of some triclopyr 

BEE moving to soil due to a rainfall event, but applications will be made to minimize the potential 

for runoff. Residues at 1 and 8 lb a.e./acre exceed the distribution of toxicity data for aquatic plants 

and exceed the effects distribution for aquatic vertebrates at the maximum labelled rate. 
 E–44 Appendix E. Herbicide Environmental Risk Assessment

Figure 5-1. Aquatic risk characterization for triclopyr BEE. 

 Residues were below all toxicity values in all soil types when considering areas that had 100 inches 

or less of annual rainfall. Areas in the United States with annual average rainfall amounts greater 

than 100 inches are restricted to coastal areas in the Pacific Northwest and a small portion of 

northwest California. These are not considered high risk areas for the introduction of ALB, thus, 

the aquatic risk from triclopyr BEE for stump treatments is very low for a majority of the country 

based on conservative estimates of exposure. Due to the significant reduction in toxicity for 

triclopyr TEA, imazapyr, and metsulfuron-methyl the potential for aquatic risk would be much less 

for these three herbicides when compared to triclopyr BEE. 

5.2.2 Terrestrial Wildlife and Domestic Animals

The risk to non-target terrestrial vertebrate wildlife and domestic animals will be negligible based 

on the proposed method of application for each herbicide and available toxicity data for mammals 

and birds. Direct application of herbicides to stumps and sprouting vegetation using hand painting 

or hand held sprayers reduces the likelihood of off-site drift and runoff. Any plausible scenarios 

where mammals and birds would be exposed to significant residues on stumps or associated 

sprouting vegetation is unlikely. Some mammals, such as white- tailed deer, may graze on sprouts 

from stumps that have been selectively treated with herbicides. However, the toxicity profiles for 

all three herbicides suggest mammals would have to consume many times their daily food 

consumption rates over an extended period of time to reach a dose that could result in any acute or 

chronic effects. Treated vegetation would become unpalatable as the sprout dies back due to 

herbicide treatment, reducing exposure time. Incidental soil ingestion by foraging mammals and 

birds would also be low because applications are directed to the stump and sprouting vegetation 

using hand held sprayers. Risk from herbicide exposure through drinking water to nontarget 

vertebrates would also be very low based on aquatic residues estimated in a small stream using 

applications of triclopyr BEE. Non-target vertebrate wildlife and domestic animals would have to 

consume many times their daily water ingestion rate to consume a dose that could result in an 

adverse effect. There is no plausible exposure scenario where that situation could occur. Risks 

from ingestion of drinking water containing triclopyr TEA, TCP, imazapyr, and metsulfuronEstimated Acute Residues (8

lb ae/ac)

Estimated Acute Residues (1 lb ae/ac)

Acute Aquatic

Acute Fish and

Acute

0

Log Triclopyr BEE
Appendix E. Herbicide Environmental Risk Assessment E–45

methyl would be even less due to their comparative lower toxicity, and in the case of imazapyr and 

metsulfuron-methyl, much lower use rates compared to triclopyr. 

5.2.3 Terrestrial Invertebrates and Plants

Herbicide applications are not expected to result in adverse impacts to terrestrial invertebrates. 

Available toxicity data for terrestrial invertebrates shows that toxicity is low for pollinators, such 

as honey bees, as well as soil borne invertebrate species for all three herbicides. The greatest 

potential for exposure would be for soil borne invertebrates however, the method of application 

suggests that any significant residues would be immediately adjacent to any treated stumps. 

All three herbicides are toxic to terrestrial plants, and non-target plants immediately adjacent to 

the treatment site would be at risk from drift and runoff. The method of application will reduce 

the risk to terrestrial plants because applications are made by hand directly to stumps and any 

associated sprouting vegetation which would minimize and drift and runoff. 

6.0 UNCERTAINTIES AND CUMULATIVE IMPACTS

The uncertainties associated with this risk evaluation arise primarily from a lack of information 

about the effects of each herbicide, its formulations, metabolites, and potential mixtures to nontarget organisms that can occur in the environment. These uncertainties are not unique to this 

assessment but are consistent with uncertainties in human health and ecological risk assessments 

with any environmental stressor. In addition, there is uncertainty in where an ALB infestation 

may occur in the United States and the extent of herbicide use in a given area because 

application is based on site-specific factors. As a way to account for some of the uncertainty 

with the available data APHIS made conservative assumptions regarding exposure and relied on 

the more sensitive effect endpoints to assess risk to human health and the environment. 

Another area of uncertainty is the potential for cumulative impacts to human health and the 

environment from the proposed use of the three herbicides together and how they may interact 

with other biotic and abiotic stressors. Areas where cumulative impacts could occur are: 1) 

repeated worker and environmental exposures to each herbicide in relation to other uses; 2) coexposure to other chemicals with a similar mode of action; and 3) exposures to other chemicals 

in mixtures and how that may affect the toxicity of each herbicide. 

From a human health perspective, the ALB program use of triclopyr and the tank mix of the 

three herbicides are expected to result in negligible cumulative impacts when considering the 

proposed use patterns. The herbicides are not expected to enter the food chain through food and 

drinking water and significantly add to triclopyr, imazapyr, and melsulfuron-methyl exposures 

and other stressors in the human population. 
 E–46 Appendix E. Herbicide Environmental Risk Assessment

Cumulative impacts may occur from the use of each herbicide as well as the three herbicides 

together in relation to other chemicals that have a similar mode of action, as well as others that 

have a different mode of action, but could result in synergistic, additive, or antagonistic effects. 

The cumulative impacts from the individual use of triclopyr or the three herbicides together are 

expected to be low because they are not persistent in the environment or bioaccumulative (Tatum 

2004). In addition, their selective use in the ALB Eradication Program and the methods of 

application reduce exposure. The low probability of exposure and favorable toxicity profile 

result in very low risk to human health and to most non-target wildlife so that any cumulative 

impacts from the use of herbicides relative to other stressors would be incrementally negligible. 

7.0 REFERENCES

BASF Corporation (BASF). 2012a. Pesticide label: Arsenal® herbicide, EPA Reg. No. 241-346. 

21 pp. 

BASF. 2012b. Safety data sheet Arsenal. Avalaible at http://www.env.gov.yk.ca/environmentyou/documents/ArsenalSafetyDataSheet.pdf, last accessed 6/25/2014. 

Boren, J.C., Lochmiller, R.L., Leslie, D.M. and D.M. Engle. 1993. Long-term effects of woody 

vegetation management on seasonal body condition of northern bobwhites. J Range Manag. 46: 

520-523. 

Brooks, J.J., Rodrigue, J.L., Cone, M.A., Miller, K.V., Chapman, B.R. and A.S. Johnson. 1995. 

Small mammal and avian communities on chemically-prepared sites in the Georgia sandhills. 

General Technical Report - Southern Research Station, USDA Forest. (No. S):21-23. 

Budavari, S. (Ed). 1989. The Merck Index: An Encyclopedia of Chemicals, Drugs, and 

Biologicals, 11th ed. Merck & Co., Inc., Rahway, New Jersey. 

Chakravarty P. and S.S. Sidhu. 1987. Effect of glyphosate hexazinone and triclopyr on in vitro 

growth of five species of ectomycorrhizal fungi. Eur J for Pathol. 17 (4-5): 204-210. 

Cox, C. 2000. Triclopyr herbicide factsheet. Journal of Pesticide Reform. 20(4): 12-19. 

Available at http://www.pesticide.org/get-the-facts/pesticide-factsheets/factsheets/triclopyr, last 

accessed 12/22/2014. 

Dow AgroSciences. 2010. Material safety data sheet Pathfinder® II herbicide. Available at 

http://www.dot.ca.gov/dist05/maint/ivm/msdsfiles/pathfinder2_msds.pdf, last accessed 

6/25/2014 

Dow AgroSciences. 2011a. Pesticide label: Garlon® 3A, EPA Reg. No. 62719-37, Alternate 

Brand Name: Element 3A. Dated October 24, 2011. 38 pp. 

Dow AgroSciences. 2011b. Pesticide label: Pathfinder® II, specialty herbicide, EPA Reg. No. 

62719-176. Dated May 6, 2011. 12 pp. 
Appendix E. Herbicide Environmental Risk Assessment E–47

Dow AgroSciences. 2011c. Material safety data sheet Garlon® 3A herbicide. Available at 

http://www.aqumix.com/pdf/Garlon3AMSDS.pdf, last accessed 6/25/2014 

DuPont. 2012a. Pesticide label: DuPontTM Escort® XP herbicide. EPA Reg. No. 352-439. Dated 

February 29, 2012. 21 pp. 

DuPont. 2012b. Material safety data sheet DuPontTM Escort® XP herbicide. Ref. 130000036195. 

Available at http://www.plunketts.net/docs/msds/escort_xp-msds.pdf, last accessed 6/25/2014. 

Estok, D., Freedman, B., and D. Boyle. 1989. Effects of the herbicides 2 4-D, glyphosate, 

hexazinone, and triclopyr on the growth of three species of ectomycorrhizal fungi. Bull Environ 

Contam Toxicol. 42(6): 835-839. 

Fowlkes, M.D., Michael, J.L., Crisman, T.L. and J.P. Prenger. 2003. Effects of the Herbicide 

Imazapyr on Benthic Macroinvertebrates in a Logged Pond Cypress Dome. Environ Toxicol 

Chem. 22(4): 900-907. 

Grisolia, C.K. 2002. A comparison between mouse and fish micronucleus test using 

cyclophosphamide, mitomycin C and various pesticides. Mutat Res. 518 :145–50. 

Hanley, T., Thompson, D.J., Palmer, A.K., Beliles, R.P. and B.A. Schwetz. 1984. Teratology and 

reproduction studies with triclopyr in the rat and rabbit. Fundam Appl Toxicol. 4:872-882. 

Harris, S.A. and K.R. Solomon. 1992. Human exposure to 2,4-D following controlled activities 

on recently sprayed turf. J Environ Sci Health. B27(1): 9-22. 

Houston, A.P.C., Visser, S. and R.A. Lautenschlager. 1998. Response of microbial processes 

and fungal community structure to vegetation management in mixed wood forest soils. Can J 

Bot. 76: 2002-2010. 

Hazardous Substances Data Bank (HSDB). 2014a. HSDB: Triclopyr, available via Toxnet at 

http://toxnet.nlm.nih.gov/cgi-bin/sis/search2/f?./temp/~TeYhGy:1, last accessed 6/27/2014. 

HSDB. 2014b. HSDB: Imazapyr, available via Toxnet at http://toxnet.nlm.nih.gov/cgibin/sis/search2/f?./temp/~JU0Z9j:1, last accessed 6/27/2014 

HSDB. 2014c. HSDB: Metsulfuron-methyl, available via Toxnet at 

http://toxnet.nlm.nih.gov/cgi-bin/sis/search2/f?./temp/~9S9p9F:1, last accessed 6/27/2014. 

Johansen, J.A. and G.H. Geen. 1990. Sublethal and acute toxicity of the ethylene glycol butyl 

ether ester formulation of triclopyr to juvenile Coho salmon (Oncorhynchus kisutch). Arch 

Environ Contam Toxicol. 19(4): 610-616. 

Kjaer, C. and U. Heimbach. 2001. Relationships between sulfonylurea herbicide treatment of 

host plants and the performance of herbivorous insects. Pest Manag Sci. 57 (12):1161-1166. 
 E–48 Appendix E. Herbicide Environmental Risk Assessment

Kreutzweiser, D.P. and S.S. Capell. 1992. A simple stream-side test system for determining 

acute lethal and behavioral effects of pesticides on aquatic insects. Environ Toxicol Chem. 11: 

993-999. 

Lajmanovich, R.C., Junges, C.M., Attademo, A.M., Peltzer, P.M., Cabagna-Zenklusen, M.C. and 

A. Basso. 2013. Individual and Mixture Toxicity of Commercial Formulations Containing 

Glyphosate, Metsulfuron-Methyl, Bispyribac-Sodium, and Picloram on Rhinella arenarum 

Tadpoles, Water Air Soil Pollut 224:1404. 

Lewer, P. and W.J. Owen. 1990. Selective action of the herbicide triclopyr. Pestic Biochem 

Physiol. 36(2): 187-200. 

Mason, R.W. and B.L. Johnson. 1987. Ergonomic factors in chemical hazard control. In: 

Handbook of Human Factors. Salveny, G; ed. John Wiley and Sons, New York, NY. pp. 772- 

741. 

Morgan, J.D., Vigers, G.A., Farrell, A.P., Janz, D.M. and J.F. Manville. 1991. Acute avoidance 

reactions and behavioral responses of juvenile rainbow trout (Oncorhynchus mykiss) to Garlon 4, 

Garlon 3A and Vision herbicides. Environ Toxicol Chem. 10(1): 73-80. 

National Pesticide Information Center (NPIC). 2002. Triclopyr general fact sheet 

http://npic.orst.edu/factsheets/triclogen.pdf last accessed 3/28/2014. 

National Research Council (NRC). 1983. Risk assessment in the Federal government: 

managing the process. National Academy Press, Washington, DC. 

Newmaster, S.G., Bell, F.W. and D.H. Vitt. 1999. The effects of glyphosate and triclopyr on 

common bryophytes and lichens in northwestern Ontario. Canadian Journal of Forest Research. 

29 (7): 1101-1111. 

Obrigawitch, T.T., Cook, G. and J. Wetherington. 1998. Assessment of effects on non-target 

plants from sulfonylurea herbicides using field approaches. Pestic Sci. 52 (3):199-217. 

Perkins, P.J., Boermans, H.J. and G.R. Stephenson. 2000. Toxicity of glyphosate and triclopyr 

using the frog embryo teratogenesis assay–Xenopus. Environ. Toxicol. Chem. 19(4): 940-945. 

Peterson, H.G., Boutin, C., Martin, P.A., Freemark, K.E., Ruecker, N.J. and M.J. Moody. 1994. 

Aquatic phytotoxicity of 23 pesticides applied at expected environmental concentrations. 

Aquatic Toxicology (Amsterdam). 28 (3-4):275-292. 

Potter, D.A., Buxton, M.C., Redmond, C.T., Patterson, C.G. and A.J. Powell. 1990. Toxicity of 

pesticides to earthworms (Oligochaeta: Lumbricidae) and effect on thatch degradation in 

Kentucky bluegrass turf. J Econ Entomol. 83: 2362-2369. 

Quaranta, A., Bellantuono, V., Cassano, G. and C. Lippe. 2009. Why amphibians are more 

sensitive than mammals to xenobiotics. PLoS ONE. 4: (11): 1-4. 
Appendix E. Herbicide Environmental Risk Assessment E–49

Schulz, C.A., Leslie, D.M. JR. Lochmiller, R.L. and D.M. Engle. 1992a. Autumn and winter 

bird populations in herbicide-treated cross timbers in Oklahoma. Am Midl Nat. 127 (2): 215-223. 

Schulz, C.A., Leslie, D.M., Lochmiller, R.L. and D.M. Engle. 1992b. Herbicide effects on cross 

timbers breeding birds. J Range Manag. 45: 407-411. 

Shackelford, D.D., Young, D.L., Mihaliak, C.A., Shurdut, B.A. and J.A. Itak. 1999. Practical 

immunochemical method for determination of 3,5, 6-trichloro-2-pyridinol in human urine: 

applications and considerations for exposure assessment. J Agric Food Chem. 47(1):177-82. 

Tatum, V.L. 2004. Toxicity, transport, and fate of forest herbicides. Wildlife Society Bulletin, 

32(4): 1042-1048. http://www.bioone.org/doi/pdf/10.2193/0091- 

7648%282004%29032%5B1042%3ATTAFOF%5D2.0.CO%3B2, last accessed 9/12/2014 

Timchalk, C., Dryzga, M.D. and P.E. Kastl. 1990. Pharmacokinetics and metabolism of 

triclopyr 3,5,6-trichloro-2-pyridinyloxyacetic acid) in Fischer 344 rats. Toxicology. 62(1): 71-87. 

Timchalk, C., Finco, D.R. and J.F. Quast. 1997. Evaluation of renal function in rhesus monkeys 

and comparison to beagle dogs following oral administration of the organic acid triclopyr (3,5,6- 

trichloro-2-pyridinyloxyacetic acid). Fundam Appl Toxicol. 36(1): 47-53. 

Timchalk, C. and R.J. Nolan. 1997. Pharmacokinetics of triclopyr (3,5,6-trichloro-2- 

pyridinyloxyacetic acid) in the beagle dog and rhesus monkey: perspective on the reduced 

capacity of dogs to excrete this organic acid relative to the rat, monkey, and human. Toxicology 

and Applied Pharmacology. 144 (2): 268-278. 

Trevathan, W. (ed.). 2002. Metsulfuron-methyl pesticide fact sheet: forest use, Oregon State 

University Extension Service, Enviormental Health Sciences Center. 2 pp. available at 

http://www.oregon.gov/odf/privateforests/docs/metsulfuronmethyl.pdf, last accessed 6/27/2014. 

United States Department of Agriculture Animal and Plant Health Inspection Service (USDA 

APHIS). 1993. Nontarget Risk Assessment for the MEDFLY Cooperative Eradication Program. 

USDA APHIS. February 1993. 

United States Department of Agricuture Forest Service (USDA FS). 1989. Final environmental 

impact statement: vegetation management in the Appalachian Mountains, Management Bulletin 

R8-MB-38, dated July, 1989. 1104 pp. 

USDA FS. 2003. Triclopyr - Human health and ecological risk assessment, Final Report. SERA 

TR 02-43-13-03b. 

USDA FS. 2005. Metsulfuron Methyl - Human health and ecological risk assessment, Final 

Report. SERA TR 04-43-17-01c, Report dated December 9, 2004, Revised on February 28, 

2005. http://www.fs.fed.us/foresthealth/pesticide/risk.shtml 
 E–50 Appendix E. Herbicide Environmental Risk Assessment

USDA FS. 2011a. Triclopyr - Human health and ecological risk assessments, Final Report. 

SERA TR-052-25-03a, Report dated May 24, 2011 with revisions as of October 20, 2011. 

http://www.fs.fed.us/foresthealth/pesticide/risk.shtml 

USDA FS. 2011b. Imazapyr – Human health and ecological risk assessment, Final Report. 

SERA TR-052-29-03a, Report dated December 16, 2011, 

http://www.fs.fed.us/foresthealth/pesticide/risk.shtml 

USDA FS. 2011c. Forest Service Program Worksheets Triclopyr TEA Formulations: 

Terrestrial Applications, Worksheet Version: 5.00.64 

USDA FS. 2011d. Forest Service Program Worksheets Triclopyr Triclopyr BEE Formulations: 

Terrestrial Applications, Worksheet Version: 5.00.64 

USDA FS. 2011e. Forest Service Program Worksheets 3,5,6-Trichloro-2-pyridinol (TCP) from 

Terrestrial Applications of Triclopyr, Worksheet Version: 6.00.01 

USDA FS. 2012. Reassessment of worker exposure rates, Final Report. SERA TR-052-30-02a, 

Report dated June 24, 2012. 

http://www.fs.fed.us/foresthealth/pesticide/pdfs/Worker_Exposure_2012.pdf 

United States Environmental Protection Agency (USEPA). 1989. Risk assessment guidance for 

Superfund volume I, human health evaluation manual (Part A). interim final. Office of 

Emergency and Remedial Response, Washington, DC. EPA/540/1-89/002 December 1989. 

USEPA. 1998. Registration eligibility document (RED). Triclopyr. EPA 738-R-98-011. 

Available at http://www.epa.gov/oppsrrd1/REDs/2710red.pdf, last accessed 6/26/2014. 

USEPA. 2002a. Triclopyr; Pesticide Tolerance. Federal Register 40 CFR 180 67: 58712 - 

58725. Available at: https://www.federalregister.gov/articles/2002/09/18/02-23746/triclopyrpesticide-tolerance, Last accessed 6/27/2014 

USEPA. 2002b. Metsulfuron Methyl; Pesticide Tolerance. Federal Register 40 CFR 180. 

67(152): 51088-51097. Available at: https://www.federalregister.gov/articles/2002/08/07/02- 

19807/metsulfuron-methyl-pesticide-tolerance, Last accessed 2/13/2015 

USEPA. 2002b. Metsulfuron Methyl; Pesticide Tolerance. Federal Register 40 CFR 180. 

67(152): 51088-51097. Available at: https://www.federalregister.gov/articles/2002/08/07/02- 

19807/metsulfuron-methyl-pesticide-tolerance, Last accessed 2/13/2015 

USEPA. 2005a. Imazapyr: Revised HED chapter of the reregistration eligibility document 

(RED). PC Code: 128821, CAS Reg 81334-34-1, Case #3078, DP Barcode: D324106. Reported 

dated: December 8, 2005. U.S. EPA E-Docket EPA-HQ-OPP-2005-0495. EPA File Name: EPAHQ-OPP-2005-0495-0004.pdf. Available at: www.regulations.gov. Last accessed 2/14/2015 
Appendix E. Herbicide Environmental Risk Assessment E–51

USEPA. 2005b. Level I screening ecological risk assessment for the reregistration of imazapyr. 

U.S. EPA E-Docket EPA-HQ-OPP-2005-0495. EPA File Name: EPA-HQ-OPP-2005-0495- 

0009.pdf. Available at: www.regulations.gov. 

USEPA. 2006. Reregistration eligibility decision for imazapyr. EPA 738-R-06-007/OPP-2005- 

0495. Available at http://www.epa.gov/oppsrrd1/REDs/imazapyr_red.pdf, last accessed 

6/27/2014. 

USEPA, 2007. Risks of imazapyr use to the Federally listed California red-legged frog (Rana 

aurora draytonii), Effects Determination for Imazapyr and the California Red-legged Frog. 

Includes Appendices A-K and Attachments 1 and 2. Documents dated July 20, 2007. Available 

at: http://www.epa.gov/espp/litstatus/effects/redlegfrog/index.html?PHPSESSID=9e2e2b07da820363425c65a13695be5a#imazapyr. Last accessed 

2/14/2015. 

USEPA. 2009. Risks of triclopyr use to Federally threatened California red-legged frog (Rana 

aurora draytonii). Including Appendices A through Q. Available at: 

http://www.epa.gov/espp/litstatus/effects/redleg-frog/. Last accessed 6/27/2014 

USPEA. 2011. Exposure factors handbook: 2011 Edition, National Center for Environmental 

Assessment, Office of Research and Development, U.S. EPA, Washington, DC. EPA/600/R090/052F, September 2011. Available at: http://www.epa.gov/ncea/efh/pdfs/efh-complete.pdf. 

Last accessed 2/14/2015. 

USEPA. 2013. Chemicals evaluated for carcinogenic potential. Office of Pesticide Programs. 

http://npic.orst.edu/chemicals_evaluated.pdf, last accessed 6/30/2014. 

USEPA. 2014a. Overview of Risk Assessment in the Pesticide Program – Human Health Risk 

Assessment, Office of Pesticide Programs 

(http://www.epa.gov/pesticides/about/overview_risk_assess.htm) last accessed 3/21/2014. 

USEPA. 2014b. Endocrine Disruptor Screening Program, Office of Chemical Safety and 

Pollution Prevention, Office of Science Coordination and Policy (http://www.epa.gov/endo/), last 

accessed 3/27/14. 

USEPA. 2014c. Ecotoxicity Categories for Terrestial and Aquatic Organisms. Technical 

Overview of Ecological Risk Assessment, Analysis Phase: Ecological Effects Characterization, 

Pesticides: Environmental Effects, 

http://www.epa.gov/oppefed1/ecorisk_ders/toera_analysis_eco.htm, last accessed 1/27/15. 

Welch, J.R., Miller, K.V., Palmer, W.E. and T.B. Harrington. 2004. Response of understory 

vegetation important to the northern bobwhite following imazapyr and mechanical treatments. 

Wildlife Soc Bull. 32: 1071-1076. 
 E–52 Appendix E. Herbicide Environmental Risk Assessment

Attachment A: Risk Estimations for Workers and General Public

USDA Forest Service Program Worksheets for triclopyr TEA and triclopyr BEE formulations for 

terrestrial applications (worksheet version: 5.00.64) and TCP from terrestrial applications of 

triclopyr (version 6.00.01) (USDA FS, 2011c,d,e) were used for the risk estimations. 

Attachment A summarizes the exposure concentrations, exposure assumptions and toxicity 

values used in the hazard quotient estimates to workers and the general public. 

 Attachment A-1 Concentrations in field solutions and exposure rates. 

Parameter/Assumption Range Value Units Reference 

Application rate 9 / 8 Lb/acre The maximum rate

per label (9 lb/acre

for TEA and TCP

per Garlon 3A label 

& 8lb/acre per

Pathfinder II label) 

Application Volume Central 25 Gal/acre USDA FS, 2011a 

Upper 40 

Lower 5 

Concentration in field solution Central 0.36/0.32 Lb/gal Calculated 

Upper 0.225/0.2

Lower 1.8/1.6

Conversion factor for lbs/gal to

mg/ml 

119.8 mg/mL per

lb/gal 

Concentration in field solution Central 4.30E+01/3.80E+01 mg/mL Calculated 

Lower 2.70E+01/2.40E+01

Upper 2.20E+02/1.90E+02

Number of Applications 1 Unitless USDA FS 2011a

Application Interval 1 Days USDA FS 2011a

Worker Exposure Rate Central 0.01 mg/kg bw USDA FS 2012

Lower 0.0001 per lb 

Upper 0.08 

Number of acres that the worker

will treat per hour Central 0.625 Acres/hr 

Lower 0.25 USDA FS 2012

Upper 1 

Number of hours of worker

exposure per day. Central 7 

Lower 6 USDA FS 2012

Upper 8 
Appendix E. Herbicide Environmental Risk Assessment E–53

Attachment A-2 Exposure assumptions for herbicide risk estimations 

Receptors Factor Value Units Reference 

Backpack Worker Absorbed dose rate USDA FS 2012 

Central 0.01 (mg/kg bw)/(lbs 

Upper 0.0001 handled per day) 

Lower 0.08 

Hours of application per day 

Central 7 Hours per day USDA FS 1989 

Upper 6 

Lower 8 

Acres treated per hour 

Central 0.625 acres/hour USDA FS 1989

Lower 0.25

Upper 1

Adult Male Body weight 80 kg USEPA 2011

Worker

Surface Area, Hands 1070 cm2 USEPA 2011 

Surface Area, Lower legs 2273 cm2 USEPA 2011 

Exposure duration 1/10/60 minutes Contaminated 

 gloves and spill,

 professional

 judgment

Adult Female Body weight 71 kg USEPA 2011 

Surface Area, Wearing shorts

and T-shirt 

Contact time/Exposure

Duration 

6473 

10 /1 

cm2

minutes/hour 

USEPA 2011 

Professional

Judgment 

Pre-teen and Body weight 56.8 kg USEPA 2011 

teenaged Child

(11-16) 

Surface Area, Wearing shorts

and T-shirt 

Contact time/Exposure

Duration 

5650 

10 / 2 

cm2

minutes/hour 

USEPA 2011 

Professional

Judgment 

Attachment A-3 Conversion factors for herbicide risk estimations. 

Receptors Factor Value Units Reference 

Conversion

Factors 

acrefoot2

gal 325900 gal/acre-foot Budavari 1989 

liters per gal 3.785 L/gal Budavari 1989

pound per milligram 453600 mg/lb Budavari 1989 

lbac2mgcm2 0.01121 mg/cm2 per lb/acre Budavari 1989

lbac2ugcm2 11.21 ug/cm2 per lb/acre Budavari 1989 

conversion for milligrams to pounds 2.204E-06 lb/mg Budavari 1989

conversion for milliliters to gallons 3785 ml/gal Budavari 1989 

Misc Liquid adhering to skin 0.008 mL/cm2 Mason and Johnson

1987 

Dislodgeable residue as proportion 0.1 none Harris and Solomon

1992
E–54 Appendix E. Herbicide Environmental Risk Assessment

Attachment A-4 Summary of toxicity values 

Compound Duration Toxicity Value

(RfD – mg/kg/day) 

Reference 

Triclopyr TEA Acute 1 

Chronic 0.05 

Triclopyr BEE Acute 1 

Chronic 0.05 

TCP Acute 0.025 

Chronic 0.012 

USDA FS, 2011a, USEPA 2002a 
Appendix F. Imidacloprid Environmental Risk Assessment F–1

F. Imidacloprid Human Health and Ecological Risk 

Assessment: Asian Longhorned Beetle 

Eradication Program 
F–2 Appendix F. Imidacloprid Environmental Risk Assessment

Table of Contents

 

EXECUTIVE SUMMARY 

1.0 INTRODUCTION 

2.1 PROBLEM FORMULATION 

2.2 Chemical Description and Product Use 

2.3 Physical and Chemical Properties 

2.4 Environmental Fate 

2.5 Hazard Identification 

2.5.1 Acute Toxicity 

2.5.2 Sub-Chronic/Chronic Toxicity 

2.5.3 Carcinogenicity/Mutagenicity 

2.5.4 Development and Reproductive Effects 

2.5.5 Endocrine Effects 

2.5.6 Immune System Effects 

2.5.7 Metabolism 

3.1 DOSE-RESPONSE ASSESSMENT 

3.2 Human Health Dose-Response Assessment 

3.3 Ecological Effects Analysis 

3.3.1 Aquatic Effects Analysis 

3.3.2 Terrestrial Effects Analysis 

3.3.3 Toxicity of formulations and metabolites to nontarget wildlife 

4.1 EXPOSURE ASSESSMENT 

4.2 Human Health Exposure Assessment 

4.2.1 Identification of Potentially Exposed Human Populations and Complete 

Exposure Pathways 

4.2.2 Exposure Evaluation 

4.3 Ecological Exposure Assessment 

4.3.1 Aquatic Exposure Assessment 

4.3.2 Terrestrial Exposure Assessment 

5.1 RISK CHARACTERIZATION 

5.2 Human Health 

5.3 Aquatic and Terrestrial Ecological Risk Characterization 

5.3.1 Aquatic 

5.3.2 Terrestrial Wildlife and Domestic Animals 

5.3.3 Terrestrial Invertebrates and Plants 

6.0 UNCERTAINTIES AND CUMULATIVE IMPACTS 

7.0 REFERENCES 
Appendix F. Imidacloprid Environmental Risk Assessment F–3

EXECUTIVE SUMMARY

The United States Department of Agriculture (USDA), Animal and Plant Health Inspection 

Service (APHIS) proposes to use the insecticide imidacloprid as part of an eradication program 

for the Asian longhorned beetle (ALB). 

USDA APHIS evaluated the potential human health and ecological risks from the proposed use 

of imidacloprid for the ALB eradication program. Risks to human health are expected to be 

negligible based on limited exposure from the proposed use pattern of imidacloprid (trunk and 

soil injection). Exposure is greatest for workers who will apply the product, but the formulation 

and the required protective equipment result in a low potential for exposure and risk. 

Ecological risks for terrestrial and aquatic non-target organisms is also expected to be low based 

on the toxicity and environmental fate of imidacloprid, and method of application. Risks to 

terrestrial invertebrates, including pollinators, is expected to be negligible based on available 

laboratory data and field collected data for pollinators that has been collected for ALB specific 

applications of imidacloprid. There is some risk to sensitive terrestrial invertebrates that consume 

vegetation from treated trees or inhabit the soil where soil injection is used. 

1.0 INTRODUCTION

This human health risk assessment (HHRA) and ecological risk assessment (ERA) provide a 

quantitative and qualitative evaluation of the potential risks and hazards to human health, 

nontarget fish, and wildlife as a result of exposure to the insecticide, imidacloprid, when used for 

controlling the ALB. 

The methods used in this HHRA to assess potential human health effects follow standard 

regulatory guidance and methodologies (NRC, 1983; USEPA, 2014a), and generally conform to 

other Federal agencies such as U.S. Environmental Protection Agency (USEPA), Office of 

Pesticide Programs (USEPA/OPP). The methods used in this ERA to assess potential ecological 

risk to nontarget fish and wildlife follow USEPA methodologies regarding eco-risk assessment, 

with an emphasis on those used by USEPA/OPP in the pesticide registration process. 

The risk assessment starts with problem formulation (identifying hazard), then the toxicity 

assessment (the dose-response assessment), and then the exposure assessment (identifying 

potentially exposed populations and determining potential exposure pathways for these 

populations). Lastly, the information from the exposure and toxicity assessments is combined to 

characterize risk (determining whether there is adverse health and eco-risk). 

2.0 PROBLEM FORMULATION

APHIS is proposing the use of the insecticide imidacloprid to treat high risk host trees that are 

not cut to prevent further spread of ALB as part of the ALB eradication. The following sections 

discuss the chemical description and product use; physical and chemical properties; 

environmental fate; and hazard identification for imidacloprid. 
F–4 Appendix F. Imidacloprid Environmental Risk Assessment

2.1 Chemical Description and Product Use

Imidacloprid (C9-H10-Cl-N5-O2) (CAS No. 138261-41-3) belongs to a class of insecticides, 

neonicotinoids, that act by binding directly to the acetylcholine binding receptor. The molecular 

structure is shown in Figure 1. Imidacloprid is manufactured in several different formulations 

that can be used for soil, seed and foliar applications. It is registered for use on a wide variety of 

agricultural commodities as well as in horticultural and turf applications and for animal health. 

Imidacloprid controls a variety of insects including sucking insects such as psyllids, aphids, 

thrips, whiteflies, rice hoppers, turf and soil insects, and some beetles. In the ALB eradication 

program, an imidacloprid formulation such as Merit® 2F will be applied through trunk or soil 

injection at the base of the tree. After application, imidacloprid is taken up and distributed 

throughout the tree (USDA APHIS, 2013). Merit® 2F (21.4% imidacloprid as an active 

ingredient and 78.6% others as inerts) contains 2 pounds of imidacloprid per gallon (Bayer 

2004). 

Figure 1 - Molecular Structure (a chloronicotinyl nitroguanidine insecticide) 

2.2 Physical and Chemical Properties

Imidacloprid is a colorless crystal with a slight odor. Its melting point is 144oC. It is not 

considered volatile due to a low vapor pressure (3 x 10-12 mmHg at 20oC). The partition 

coefficient (log Kow) is 0.57 at 21 o

C suggesting it would not bioaccumulate in lipids. The 

solubility in water is 0.61 grams per liter (g/L) at 20°C. The estimated Henry’s constant value is 

2 x 10-15 atm-m3/mole at 20oC. 

2.3 Environmental Fate

The environmental fate describes the processes by which imidacloprid moves and transforms in 

the environment. The environmental fate processes include: 1) mobility, persistence, and 

degradation in soil, 2) movement to air, 3) migration potential to groundwater and surface water, 

and 4) plant uptake. 

In soil, imidacloprid soil organic carbon-water partitioning coefficient (Koc)1 values range from 

157 to 810 in various soil types and organic carbon levels suggesting that in most cases 

imidacloprid does not adsorb strongly to soil particles (USDA FS, 2005; HSDB 2014). The 

adsorption capacity of imidacloprid to soil is correlated with the amount of soil organic matter 

(Liu et al., 2006). Imidacloprid is less mobile in organic carbon and clay soils and more mobile 

in sand and gravel soils (HSDB, 2014). The leaching potential also depends on the soil type and 

organic content level in soil. Several studies have shown that imidacloprid does not leach 

1 KOC values are useful in predicting the mobility of organic soil contaminants. Higher KOC values 

correlate to less mobile organic chemicals while lower KOC values correlate to more mobile organic 

chemicals.
Appendix F. Imidacloprid Environmental Risk Assessment F–5

to ground water in silty loam and silt soils with high organic matter content.(Miles, Inc., 1992; 

Rouchaud et al., 1994). A leaching study in fine sandy loam soil demonstrated the potential for 

imidacloprid residues to move downward through the soil with percolating water (Felsot et al., 

1998). 

Soil photolysis half-lives range from 19 to 39 days (CA DPR, 2006a; Graebing and Chib, 2004). 

Aerobic soil metabolism half-lives show imidacloprid to be persistent with half-lives ranging 

from 83 days to greater than one year (CCME, 2007). Imidacloprid degrades more rapidly in 

soil under aerobic conditions with vegetation (half-lives of 48 days) compared to soil without 

vegetation (half-lives of 190 days) (HSDB, 2014). Imidacloprid degrades under anaerobic 

aquatic conditions more rapidly than under aerobic soil conditions (HSDB, 2014). At the soil 

surface, imidacloprid has a photolysis half-life of 39 days. When incorporated into soil, the 

photolysis dissipation half-life of imidacloprid varies from 27 to 229 days (CA DPR, 2006a). 

The half-lives tend to increase with increasing pH values but decrease in the presence of light 

(Sarkar et al., 2001; CA DPR, 2006a). Soil half-life values in the absence of light are as high as 

229 days in field studies and 997 days under laboratory conditions (CA DPR, 2006a). Several 

field dissipation studies have been conducted with imidacloprid under various crop, soil, and 

weather conditions with a resulting half-life range of 7 to 107 days (USDA FS, 2005). 

Imidacloprid degrades under anaerobic aquatic conditions more rapidly than under aerobic soil 

conditions (HSDB, 2014). 

In air, imidacloprid is expected to exist solely in the particulate phase and not expected to 

volatilize into the ambient atmosphere based on the low reported volatility (3 x 10-12 mm Hg) 

and Henry’s Law Constant2 (2 x 10-15 atm m3/mole) (HSDB, 2014). 

Imidacloprid is considered soluble in water with values ranging from 510 to 610 mg/L (USDA 

FS, 2005). In aquatic environments imidacloprid is resistant to hydrolysis at all relevant pH 

values but is susceptible to aquaeous photolysis with half-lives of less than 5 hours (USDA FS, 

2005). 

In plants, imidacloprid applied through trunk or soil injection to host trees will translocate 

throughout the tree. These applicaton methods were used previously in the ALB eradication 

programs in New York and Illinois. During the environmental monitoring, paired leaf and twig 

samples from eight different species of host trees in New York were collected periodically for 15 

months to test for the presence of imidacloprid residues. Tree species included Norway, 

sycamore, sugar and silver maple, poplar, elm, hackberry and mountain ash. Plant tissue 

sampling of trees treated with imidacloprid using the same application methods indicated that 

both leaves and twigs have detectable residues of imidacloprid concentrations (USDA APHIS, 

2002). The detected concentrations in leaves were higher than the detected concentrations in 

twigs. Imidacloprid appears to be distributed predominately in areas that are rapidly growing 

(i.e., leaves) based on the sampling data. The sampling data also shows that uptake and 

distribution varied by tree species. Blossom sampling results (USDA APHIS, 2001; 2002) show 

that imidacloprid residues were either not detected or detected at low levels not 

2 Henry’s law constant is the ratio of a chemical’s concentration in the air to its concentration in water at equilibrium.
F–6 Appendix F. Imidacloprid Environmental Risk Assessment

Quantifiable (0.03 to 0.099 ppm) except for one blossom sample that contained 0.13 ppm (USDA 

APHIS, 2002). 

The half-life of imidacloprid on vegetation from field dissipation studies has been shown to 

range from 1.17 days in potatoes to 9.8 days on turf. Because applications in this program are 

not foliar, this route of degradation is not expected to be significant since imidacloprid will be 

translocated by plant roots into the tree or distributed throughout the tree from trunk injection 

sites by the tree’s vascular system. The main breakdown products of imidacloprid in plants 

includes a monohydroxy metabolite, imidacloprid guanidine, imidacloprid olefin, and a 

monoglucoside of 6-chloropicolyl alcohol (CA DPR, 2006a). 

2.4 Hazard Identification

Imidacloprid, a neonicotinoid insecticide, results in toxicity through binding or partial binding to 

specific areas of the nicotinic acetylcholine receptor (nAChR). Acetylcholine is an important 

neurotransmitter in most animals. It is released at the nerve synapse in response to a membrane 

depolarization that is the initiator of nerve transmission. Acting as an inhibitor at nicotinic 

acetylcholine receptors, imidacloprid disrupts the nervous system. Imidacloprid is more toxic to 

insects than to mammals because it has higher binding strength to the nicotinic receptors of 

insect nerve cells than to mammalian receptors (Tomizawa and Casida, 2003; Gervais et al., 

2010). 

Imidacloprid has moderate acute oral toxicity and low acute dermal and inhalation toxicity, but is 

not an eye or dermal irritant, or dermal sensitizer. The primary target of imidacloprid is the 

nervous system. Imidacloprid has been shown to have thyroid and/or liver effects in chronic 

dietary studies in rats and dogs. These reported impacts occurred at levels above those 

anticipated from imidacloprid use in the ALB eradication program. Reports of imidacloprid 

deliberate self-poisoning or accidental exposure in humans have shown low lethality (Mohamed 

et al., 2009; Gervais et al., 2010; Lin, et al., 2013; Kumar et al., 2013). Signs of toxicity include 

drowsiness, dizziness, vomiting, disorientation, and fever (Gervais et al., 2010). Other more 

serious health effects from exposure to imidacloprid include compromised respiratory, 

cardiovascular, and neurological function such as dyspnoea/apnoea, coma, tachycardia, 

hypotension, mydriasis, and bradycardia (Lin et al., 2013). Kumar et al. (2013) reported signs of 

acute inhalation intoxication from an accidental imidacloprid exposure that included severe 

gastrointestinal symptoms, respiratory distress, and neuropsychiatric features. Signs of toxicity 

from accidental inhalation of a pesticide containing 17.8% imidacloprid included becoming 

disoriented, agitated, incoherent, with sweating and breathlessness (Gervais et al., 2010). Two 

fatal incidents of deliberate ingestion were associated with suicides (Gervais et al., 2010; 

Iyyadurai et al., 2010). 

2.4.1 Acute Toxicity

 

Multiple acute toxicity studies have been conducted using imidacloprid on several 

mammalian species. These studies along with other available data have been summarized in 

several reports (CA DPR, 2006b; USDA FS, 2005; USEPA, 2005; USEPA, 2010). Acute 

oral median lethal toxicity values range from 131 mg/kg for the mouse to 475 mg/kg for the 

rat, suggesting moderate acute toxicity to mammals. Inhalation and dermal toxicity is 
Appendix F. Imidacloprid Environmental Risk Assessment F–7

considered low for imidacloprid with LC50 and LD50 values greater than 5.33 mg/L and 

2,000 mg/kg, respectively. Acute oral and inhalation sublethal effects have been measured in 

the rat and mouse with oral no observable effect levels (NOEL) ranging from 10 to 50 

mg/kg/day and inhalation NOELs ranging from 3.4 to 192 mg/kg/day (CA DPR, 2006b). 

Sublethal impacts noted in these studies were apathy, labored breathing, trembling and 

staggering gait. An acute neurotoxicity study in rats reported a lowest observable adverse 

effect level (LOAEL) of 42 mg/kg (USEPA, 2010). The California Department of Pesticide 

Regulations (CA DPR) has used this value to set the acute benchmark dose (BMD0.5) of 9 

mg/kg/day while USEPA determined that the reported NOEL was a LOEL and calculated an 

acute Population-Adjusted Dose (aPAD)/Reference Dose (RfD) of 0.14 mg/kg/day based on 

the application of uncertainty factors (CA DPR, 2006b; USEPA, 2010). An uncertainty 

factor of 300 (10X for interspecies extrapolation, 10X for intraspecies variation, and 3X for 

the use of a LOAEL due to the lack of a NOAEL) was applied to the aPAD (USEPA, 2010). 

The acute oral LD50 values obtained from rat studies for one of the available imidacloprid 

formulations, Merit® 2F, demonstrates lower acute toxicity when compared to the technical 

active ingredient (table 2-1). The acute inhalation toxicities are very low (LC50 > 2.0 mg/L) 

for both the technical imidacloprid and the Merit® 2F formulation. 

 

Table 2-1. Comparative acute mammalian toxicity between the technical active ingredient 

and the formulated imidacloprid . 

* Source: USDA FS, 2005; Bayer, 2008 

 

2.4.2 Sub-Chronic/Chronic Toxicity

Chronic toxicity is a condition caused by repeated or long-term exposure to low doses of a 

toxic substance. Several sub-chronic studies have been submitted to support the registration 

of imidacloprid. Studies range in duration from 21 to 107 days testing under oral, dermal, 

and inhalation exposures (CA DPR, 2006b; USDA FS, 2005) (table 2-2). 

Toxicity Study Technical Imidacloprid Merit® 2F* 

 

Acute Oral LD50

(rat) 

424 mg/kg (M)/450 mg/kg 

(F) 

>4,870 mg/kg (M)/4,143 

mg/kg (F) 

Acute Dermal 

LD50 

>5,000 mg/kg (rat) > 2,000 mg/kg (rabbit) 

Acute Inhalation 

LC50 (rat) 

>5.3 mg/L >5.33 mg/L 

Dermal Irritation 

(rabbit) 

No irritation No irritation 

Ocular (rabbit) No irritation Mild Irritation 

(Minimally irritating) 

Sensitization 

(Guinea pig) 

Non-sensitizing Non-sensitizing 

M= male; F=Female 
F–8 Appendix F. Imidacloprid Environmental Risk Assessment

Table 2-2. Subchronic mammalian effects studies using imidacloprid. 

Test Animal Exposure/Duration NOEL

(mg/kg/day)

LOEL (mg/kg/day)

Rat Oral/98 days 14.0 61.0 

Rat Inhalation/28 days 0.9 5.2 

Rat Oral Neurtoxicity/91 days 9.3* NR 

Mouse Oral/107 days 86.0 427.0 

Dog Oral/28 days 7.3 31.0 

Dog Oral/91 days 8.0 24.0 

Rabbit Dermal/21 days >1,000 >1,000 

*Value represents the lowest reported NOEL from the study; NR = Not Reported

Chronic studies have been conducted using the rat, dog and mouse in oral exposures ranging 

from one to two years. The rat was the most sensitive test species with a NOEL and LOEL of 

5.7 and 17 mg/kg/day, respectively. Comparative values for the dog and mouse were higher 

with NOELs of 15 and 47, respectively, and LOELs of 41 and 143 mg/kg/day (CA DPR, 

2006b). The NOEL reported in the two year rat study was used by USEPA/OPP to set the 

chronic reference dose (RfD) of 0.057 mg/kd/day (USEPA, 2010). RfD is “an estimate of daily 

oral exposure to the human population that is likely to be without an appreciable risk of 

deleterious effects during a lifetime” (USEPA Risk Assessment Glossary, 

http://www.epa.gov/risk_assessment/glossary.htm). 

2.4.3 Carcinogenicity/Mutagenicity

There is no evidence of carcinogenic potential in the rat and mouse carcinogenicity studies, and 

there is no concern for mutagenicity. Based on studies with rats and mice, the USEPA has 

classified imidacloprid into Group E, no evidence of carcinogenicity (USEPA, 2005; 2013). A 

study of human lymphocytes exposed to greater than 5,200 μg/mL of imidacloprid 

demonstrated a slight increase in chromosome abnormalities in vitro, but this result was not 

found with in vivo tests (Gervais et al., 2010). Imidacloprid is not considered mutagenic or 

genotoxic based on the weight of evidence from several in vitro and in vivo studies. The results 

of an in vitro study on human peripheral blood lymphocytes indicated that imidacloprid at 

concentrations less than 20 microMolar is not genotoxic to human lymphocytes (Costa et al., 

2009). 

2.4.4 Development and Reproductive Effects

A two generation reproductive study using the rat resulted in a NOEL of 13 mg/kg/day for 

adult rats based on a decrease in premating body weights. The offspring NOEL was also 13 

mg/kg/day based on decreased pup body weight in both litters (CA DPR, 2006b). 

Developmental studies using the rat reported a maternal and developmental NOEL of 13 

mg/kg/day, and in a rabbit study, the maternal and developmental NOEL was reported as 24 

mg/kg/day. A chronic fertility study showed that exposure to imidacloprid at doses of 225 and 

112 mg/kg for 60 days can cause histological damage to testicular tissue, sperm mortality, and 

decreased testosterone levels in mature male rats (Najafi et al., 2010). 
Appendix F. Imidacloprid Environmental Risk Assessment F–9

2.4.5 Endocrine Effects

A literature search did not identify any study indicating the potential of imidacloprid to affect 

the endocrine system. Imidacloprid is among the group of 58 pesticide active ingredients on 

the initial list to be screened under the USEPA Endocrine Disruptor Screening Program. 

However, the list of chemicals was generated based on exposure potential, not based on 

whether the pesticide is a known or likely potential endocrine disruptor (USEPA, 2014b). 

Imidacloprid is not among the European Union (EU) list of chemicals with the potential to 

impact the endocrine system. The EU list includes three categories: Category 1 – endocrinal 

effect recorded at least on one type of animal; Category 2 – a record of biological activity in 

vitro leading to disruption; and Category 3 – not enough evidence or no evidence data to 

confirm/disconfirm endocrinal effect of tested chemicals (Hrouzková and Matisova, 2012). 

2.4.6 Immune System Effects

A literature review of earlier studies indicates that imidacloprid does not have a direct effect 

on the immune system in mammals (USDA FS, 2005). Recent studies suggest that exposure 

to imidacloprid may induce immunotoxicity (Mohany et al., 2011; Gawade et al., 2013; 

Badgujar et al., 2013). A study in male albino rats indicated that exposure to 1/100 of the 

LD50 for imidacloprid can induce immunotoxicity, oxidative stress, lipid peroxidation, and 

hepatotoxicity (Mohany et al., 2011). A developmental immunotoxicity study in Wistar rats 

(Gawade et al., 2013) indicated that continuous exposure to imidacloprid during development 

adversely affected the immune system. An oral exposure study in BALB/c mice over 28 

days showed that imidacloprid has immunosuppressive effects at doses >5 mg/kg (Badgujar 

et al., 2013). 

2.4.7 Metabolism

The imidacloprid molecule (composed of a pyridinyl moiety and an imidazolidine ring) is 

metabolized through two main routes of metabosim. The first route is oxidative cleavage 

that frees the pyridinyl moiety to yield 6-chloronicotinic acid. Then, 6-chloronicotinic acid is 

mostly conjugated with glycine to form a hippuric acid-type conjugate. Minor quantities of 

6-chloronicotinic acid are dechlorinated to form methylmercaptonicotinic acid and 

derivatives. The second route is hydroxylation of the imidazolidine ring to form 4- or 5- 

hydroxyl imidacloprid (HSDB, 2014). In mammalian studies imidacloprid has been shown 

to be rapidly excreted in feces and urine. In an oral dosing rat study using radiolabelled 

imidacloprid approximately 90 percent of the radioactivity was excreted through the urine 

and feces within 24 hours, with 96 percent excreted after 48 hours (USDA FS, 2005). 

3.0 DOSE-RESPONSE ASSESSMENT

3.1 Human Health Dose-Response Assessment

A dose-response assessment evaluates the dose levels (toxicity criteria) for potential human 

health effects including acute and chronic toxicities. 
F–10 Appendix F. Imidacloprid Environmental Risk Assessment

USEPA derived an acute reference dose (RfD) of 0.14 mg/kg for imidacloprid based on the 

LOAEL of 42 mg/kg body weight divided by an uncertainty factor of 300 (10 for interspecies 

extrapolation, 10 for intraspecies sensitivity, 3 for using a LOAEL to approximate a NOAEL) 

(USEPA, 2010). 

USEPA also derived a chronic RfD of 0.057 mg/kg/day based on chronic toxicity and 

carcinogenicity studies using the rat. The NOAEL was estimated to be 5.7 mg/kg/day and the 

LOAEL was set at 16.9 mg/kg/day based on increased occurrence of mineralized particles in the 

thyroid gland of male rats (USEPA, 2010). 

USEPA has not established a maximum contaminant level for imidacloprid in water (USEPA, 

2014c). USEPA developed a chronic drinking water level of comparison (DWLOC) (1,755 ppb) 

for imidacloprid for an aggregate dietary exposure analysis during registration. The DWLOC 

determines the theoretical upper limits for a pesticide’s concentration in drinking water (USEPA, 

2005) 

For the ALB eradication environmental monitoring program in Suffolk County, New York, the 

New York State "Imidacloprid Groundwater Monitoring Project Plan” established an "action 

threshold" of 25 ppb (half of the New York State drinking water standard) for imidacloprid 

(USDA APHIS, 2007). 

3.2 Ecological Effects Analysis

This section of the risk assessment discusses available ecological effects data for terrestrial and 

aquatic biota. Available acute and chronic toxicity data are summarized for all major taxa and 

will be integrated with the exposure analysis section to characterize the risk of imidacloprid to 

nontarget wildlife and domestic animals. Information in this section was gathered from on-line 

databases and searches for relevant peer reviewed and non-peer reviewed literature. In cases 

where multiple toxicity values were located for the same test species, the lowest value was 

generally used in the effects analysis. This was particularly the case for aquatic invertebrate 

toxicity data where multiple acute lethality values were available for the same test species. Other 

values that were not selected were within the range of effects data that was reviewed during the 

literature search. 

3.2.1 Aquatic Effects Analysis

3.2.1.1 Fish and Amphibians

Imidacloprid acute toxicity to fish and amphibians is low based on the available acute 

median lethal concentrations (USEPA, 2008; USDA FS, 2005; Feng et al., 2004; Jemec 

et al., 2007) (table 3-1). Values typically exceed 100 mg/L suggesting that imidacloprid is 

practically non-toxic to fish and amphibians. Sublethal toxicity based on available no 

observable effect concentrations (NOEC) data ranges from 25 to 58 mg/L for fish with 

effects such as erratic swimming behavior, discoloration, quiescence and labored 

respiration noted at higher concentrations (USDA FS, 2005). 

 
Appendix F. Imidacloprid Environmental Risk Assessment F–11

Table 3-1. Acute aquatic vertebrate toxic ity data for imidacloprid. 

Test Organism Endpoint/Length Toxicity Value (mg/L) NOEC (mg/L) 

Bluegill 96-hour LC50 >105 25 

Rainbow Trout 96-hour LC50 229.1 52.1 

Zebrafish 96-hour LC50 241.0 NR 

Sheepshead Minnow 96-hour LC50 163 58.2 

Rana limnocharis 48-hour LC50 165 30 

Rana hallowell 48-hour LC50 219 101.2 

Hypsiboas pulchellus 96-hour LC50 84.91 NR 

NR = not reported 

Limited chronic fish toxicity data is available; however, in a 98-day flow-through 

exposure the rainbow trout LOEC and NOEC were 2.3 and 1.2 mg/L, respectively, based 

on a statistically significant reduction in length (USDA FS, 2005). Sanchez-Bayou and 

Goka (2005) dosed rice paddies at a rate 1.5 times the proposed rate and measured 

several sublethal responses in Japanese medaka over an approximate three month period. 

No statistical analysis was completed; although, the authors noted a lack of 

malformations in imidacloprid treated fields, and did note an increase in parasitism rates 

by the ectoparasite, Trichodenia, when compared to controls. Measured concentrations 

of imidacloprid ranged from 239.2 μg/L immediately after dosing to 1.1 μg/L 118 days 

post treatment. Ruiz de Arcautea et al. (2014) reported various genotoxic effects to 

tadpoles of the Montevideo tree frog, Hypsiboas pulchellus, at concentrations ranging 

from 15 to 45 mg/L in 48 and 96 hour exposures. 

 3.2.1.2 Aquatic Invertebrates 

Aquatic invertebrates are more sensitive to imidacloprid when compared to fish or 

amphibans with acute median toxicity values in the high part per trillion range to greater 

than 100 mg/L depending on the test species (Song et al., 1997; Song and Brown, 1998; 

USDA APHIS, 2002; USEPA, 2008; Overmyer et al., 2005; USDA FS, 2005; Paul et al., 

2006; Key et al., 2007; Alexander et al., 2007; Stoughton et al., 2008; Roessink et al. 

2013) (table 3-2). 

Sublethal effects have also been observed for various aquatic invertebrates under acute 

and chronic exposures. Agatz et al. (2014) reported feeding inhibition for Gammarus 

pulex in four day exposures to imidacloprid residues as low as 30 μg/L. Agatz et al. 

(2013) reported similar effects to the cladoceran, D. magna, where 50% reductions in 

feeding occurred at an exposure concentration of 1.83 mg/L. Reduced feeding resulted in 

impacts to survival, growth, and reproduction; however, the presence of surplus food 

after exposure allowed for recovery. In cases of limited food availability, recovery in 

growth was not observed. Alexander et al. (2007) reported feeding inhibition for the 

mayfly, Epeorus longimanus, and the aquatic oligachaete, Lumbriculus variegatus, in 

pulsed 24-hour doses using imidacloprid at concentrations above 0.5 μg/L. Survivorship 

effects were noted at concentrations between 1 and 10 μg/L. 

 
F–12 Appendix F. Imidacloprid Environmental Risk Assessment

Table 3-2. Representative acute aquatic invertebrate toxicity data for imidacloprid. 

Test Organism Endpoint/Length Toxicity Value (μg/L) NOEC(μg/L)

Epeorus longimanus 96-hour LC50 0.65 NR 

Caenis horaria 96-hour EC50 1.77 NR 

Cloeon dipterum 96-hour EC50 1.02 NR 

Daphnia magna 48-hour EC50 85200 42000 

Daphnia magna 48-hour EC50 10440 NR 

Chironomus tentans 96-hour LC50 5.75 1.03 

C. tentans 10-day LC50 3.17 0.67 

Hyallela azteca 96-hour LC50 65.43 54.24 

H. Azteca 96-hour LC50 526.0 0.35 

Simulium vittatum 48-hour LC50 6.75 – 9.54 NR 

Lumbriculus variegatus 96-hour EC50 6.2 NR 

Aedes aegypti (4th instar) 72-hour LC50 84.0 NR 

Aedes aegypti (adult) 72-hour LC50 >6300 NR 

A. taeniorhynchus 48-hour LC50 13.0 NR 

Mysidopsis bahia 96-hour LC50 38.0 32.0 

Palaemonetes pugio 96-hour LC50 308.8 100 

Artemia sp. 48-hour LC50 361230 NR 

Crassostrea virginica 96-hour EC50 >100000 145000 

NR = not reported 

 

Chronic toxicity to aquatic invertebrates has been assessed in both marine and 

freshwater invertebrate species. In two flow-through studies using M. bahia, the 

maximum allowable toxicant concentration (MATC) values for reproductive success 

ranged from greater than 643 ng/L to 849 ng/L while the MATC for growth ranged 

from 230 to 3806 ng/L (USDA FS, 2005). Stoughton et al. (2008) conducted 28-day 

continuous and pulse dose studies using the midge, C. tentans, and the freshwater 

amphipod, H. azteca. The C. tentans NOECs for survival and weight were both 1.14 

μg/L under continuous exposure and 3.47 μg/L under a pulse dose. H. azteca was 

comparatively less sensitive with a NOEC for survival and weight of 3.44 and 11.46 

μg/L, respectively, under continuous exposure, and a NOEC of 3.53 and 11.93 μg/L for 

the same endpoints under a pulse dose exposure. Chronic toxicity testing using the less 

sensitive D. magna resulted in NOEC and LOEC values of 1.8 and 3.6 mg/L, 

respectively (USEPA, 2008). Pestana et al. (2009) reported reduced leaf litter 

decomposition related to decreased feeding activity by stoneflies after 

pulsed dose exposures in mesocosms receiving three doses of imidacloprid at 17.60 

μg/L. No effects were observed at the next lowest concentration, 1.63 μg/L. Nyman et 

al. (2013) reported reduced feeding activity and lipid content in 14- and 21-day 

exposures to the amphipod Gammarus pulex exposed to imidacloprid residues as low 

as 15 μg/L. 

Sublethal impacts to some aquatic invertebrates that feed on leaf litter containing 

imidacloprid have been observed, as well as effects on decomposition rates 

(Kreutzweiser et al., 2007; Kreutzweiser et al., 2008; Kreutzweiser et al., 2009). 

Mortality to leaf-shredding insects occurred at higher rates that were intentionally 

overdosed; however, significant mortality did not occur to shredding insects such as 
Appendix F. Imidacloprid Environmental Risk Assessment F–13

Pternarcys dorsata and Tipula sp. at typical field applications. Feeding inhibition was 

observed at imidacloprid leaf concentrations of 18-30 μg/g which have been observed 

in the field under normal applications (Kreutzweiser et al., 2009). 

In a review of a microcosm study that was submitted to support registration of 

imidacloprid negative impacts to some blue green algae, copepods, and several insect 

groups were observed at measured average doses ranging from 19 to 180 μg/L (USDA 

FS, 2005). The NOEC based on total number of invertebrates and species richness was 

determined to be 2 μg/L from the study where tanks were dosed biweekly four times. 

3.2.1.3 Aquatic Plants 

Aquatic plant studies testing technical and formulated material on blue green algae, 

freshwater diatoms, and green algae have demonstrated low toxicity (USDA FS, 2005; 

USEPA, 2008) (table 3-3). 

Table 3-3. Acute aquatic plant toxicity data for imidacloprid. 

Test Organism Endpoint/Length Toxicity Value (mg/L) NOEC (mg/L)

Anabaena flosaquae* 120-hour EC50 32.8 24.9

Navicula 

pelliculos

120-hour EC50 NR 6.69 

Scenedesmus subspicatus 96-hour EC50 >10 10 

Selanastrum capricornutum 120-hour EC50 >119 >119 

*Formulated material; NR = Not reported 

3.2.2 Terrestrial Effects Analysis

 

3.2.2.1 Mammals and Birds

Imidacloprid has moderate acute oral toxicity to wild mammals based on the available 

toxicity data used to evaluate human health effects. Imidacloprid is considered toxic to 

birds with acute oral median toxicity values ranging from 41 to 152 mg/kg (USDA 

APHIS, 2002a; USEPA, 2008; USDA FS, 2005) (table 3-4). Reproduction studies using 

the mallard and bobwhite quail have shown NOECs of approximately 125 ppm for both 

species. 
F–14 Appendix F. Imidacloprid Environmental Risk Assessment

Table 3-4. Acute avian oral and dietary toxicity data for imidacloprid. 

Test Species Study LD50/LC50 NOEL/LOEL 

House Sparrow Acute Oral 41.0 3.0/NR 

Pigeon Acute Oral NR NR/12.5 

Canary Acute Oral NR 10.0/NR 

Ringed Turtle Dove Subacute Dietary NR NR/228.0 

Bobwhite Quail Acute Oral 152.3 25.0 

Subacute Dietary 1535.87 69.0 

Chronic Reproductive NA 126.0 

Mallard Subacute Dietary 4797.0 69.0 

Chronic Reproductive NA 125.0 

NR = not reported; NA = not available 

 

3.2.2.2 Reptiles and Amphibians (terrestrial phase)

No acute or chronic toxicity data appears to be available for reptiles or terrestrial phase 

amphibians based on a review of the literature and databases containing toxicity data for 

imidacloprid. Available data for the aquatic phase of amphibians demonstrates low 

toxicity and is comparable to the data for surrogate fish species. In cases where effects 

data are lacking for reptiles USEPA assumes that the sensitivity is comparable to birds 

which would suggest that imidacloprid is toxic to reptiles. There is uncertainty in that 

assumption due to physiological and life history differences between birds and reptiles. 

3.2.2.3 Terrestrial Invertebrates and Soil Microorganisms

A substantial amount of imidacloprid toxicity data has been collected for a variety of 

insects and other arthropods with results varying based on the terrestrial invertebrate 

species tested, the method of application, and other environmental factors (USDA FS, 

2005; Pisa et al., 2015). Most of this data has been collected to assess potential impacts 

to parasitoids and predators to determine compatability of imidacloprid in integrated pest 

management. In general, imidacloprid offers reduced effects to these types of 

invertebrates compared to broad spectrum insecticides. Impacts to susceptible insects 

that feed on treated trees are expected, but due to the method of application and the 

treatment of specific ALB host trees, the effects are expected to be localized and not 

widespread. 

Due to concerns with pollinators such as honey bees several toxicity studies have been 

conducted assessing acute and chronic effects of imidacloprid to these species. Technical 

and formulated imidacloprid is considered acutely toxic to honey bees and other related 

bee species by oral and contact exposure. Median lethal toxicity values range from 3.7 to 

230 ng/bee (Schmuck et al., 2001; Tasei, 2002; USDA FS, 2005; USEPA, 2008). Acute 

sublethal effects in laboratory studies have shown that the NOEC may be less than 1 

ng/bee (USDA FS, 2005). Imidacloprid metabolite toxicity to honey bees is variable 

with some of the metabolites having equal toxicity to imidacloprid while other 

metabolites are considered practically non-toxic (USDA FS, 2005). 

Due to concerns regarding the potential sublethal impact of imidacloprid to honey bees, 
Appendix F. Imidacloprid Environmental Risk Assessment F–15

several studies have been conducted to determine potential effects in laboratory and field 

situations. Studies to assess the effects of imidacloprid on homing behavior, colony 

development, foraging activity, reproduction, wax/comb production, colony health, as 

well as other endpoints, reveal conflicting results with similar effects not typically 

observed at test concentrations under realistic exposure scenarios (Tasei et al., 2000; 

Tasei et al. 2001; Tasei, 2002; Bortolloti et al., 2003; Maus et al., 2003; Morandin and 

Winston, 2003; Stadler et al, 2003; Schmuck, 2004; Johnson, 2012; Fischer et al., 2014; 

Scholer and Krishik, 2014; 

Recent studies have reported mixed results regarding impacts to pollinators at field 

relevant doses. Feltham et al. (2014) reported a reduction in pollen collection in the 

field for bumblebees exposed in the laboratory to sugar water and pollen imidacloprid 

concentrations of 0.7 and 6.0 μg/L, respectively. No differences were reported for nectar 

collection efficiency between controls and imidacloprid exposed bees. 

Imidacloprid lethality and sublethal studies have been conducted with various earthworm 

species. Lethality varies with LC50 values ranging from 1.5 to 25.5 ppm (Pisa et al., 

2015). Sublethal effects have been noted in various earthworm species as well. Effects 

on reproduction and behavior, such as burrowing, have been noted in various species and 

under various exposure conditions with effects seen at concentration ranging from 0.1 

ppm and above (Bhattacharya and Sahyu, 2013; Pisa et al., 2015). Earthworm studies 

testing the effects of imidacloprid applied for the ALB have reported effects in both soil 

and trunk injections. Kreutzweiser et al. (2009) reported no mortality to earthworms that 

were fed scenescent leaves from trunk injected maple trees containing 3 to 11 mg/kg of 

imidacloprid. However they did report sublethal impacts such as antifeeding behavior 

and earthworm weight loss in addition to reduced leaf litter degradation. In soil treatment 

studies using imidacloprid applied for ALB control a LC10 and LC50 of 2 and 5.7 mg/kg, 

respectively, were reported for, the earthworm Dendrobaena octaedra. Sublethal effects 

such as weight loss were reported at 3 mg/kg. The other earthworm species tested, 

Eisenia fetida, was less sensitive with a reported LC50 of 25 mg/kg and sublethal effects 

noted at 14 mg/kg (Kreutzweiser et al., 2008). 

3.2.2.4 Terrestrial Plants

Toxicity testing using terrestrial plants is not typically required in the registration of 

insecticides and little data exists regarding potential effects. Of the available data, no 

impacts to treated terrestrial plants have been noted in forestry or agricultural settings 

(Westwood et al., 1998; USDA FS, 2005). 
F–16 Appendix F. Imidacloprid Environmental Risk Assessment

3.2.3 Toxicity of formulations and metabolites to nontarget wildlife

The aquatic toxicity of the primary metabolites of imidacloprid in the environment which are 

urea-based metabolites, designated as NTN 33823 and NTN 33519, and 6-chloronicotinic 

acid have been evaluated for H. azteca and C. tentans which represent the more sensitive 

aquatic species. Based on the range of toxicity values for each of the metabolites compared 

to imidacloprid, the toxicity for each metabolite is several orders of magnitude less than the 

parent (USDA FS, 2005) (table 3-5). 

 

Table 3-5. Comparative aquatic toxicity data for imidacloprid and associated metabolites. 

Test Species Chemical Endpoint/Length Toxicity Value (μg/L) NOEC

(μg/L) 

Hyallela azteca imidacloprid 96-hour LC50 526.0 0.35

NTN 33823 96-hour LC50 51800 22100

NTN 33519 96-hour LC50 >94830 94830

C. tentans imidacloprid 96-hour LC50 5.75 1.03

NTN 33823 96-hour LC50 >82800 8190

NTN 33519 96-hour LC50 >99800 99800

6-chloronicotinic acid 96-hour LC50 >1000 1000

Formulation aquatic toxicity data does not appear to be available for Merit®, or some of the 

other available formulations for use against ALB; however, aquatic toxicity data with other 

formulations containing imidacloprid have demonstrated comparable, or increased aquatic 

toxicity in studies conducted with the technical material (Stoughton et al., 2008). The 

complete list of inerts is considered confidential business information and therefore is 

currently not available for review for most formulations. The material safety data sheet for 

Merit® states that glycerin is present although the exact quantity is unknown. 

 

4.0 EXPOSURE ASSESSMENT

4.1 Human Health Exposure Assessment

Exposure assessments estimate the potential exposure of humans to imidacloprid. The exposure 

assessment begins with the use and application method for imidacloprid. An identified exposure 

pathway for imidacloprid includes (1) a release from an imidacloprid source, (2) an exposure 

point where contact can occur, and (3) an exposure route such as ingestion, inhalation, or dermal 

contact (USEPA 1989). Exposures for the various subgroups of the population are qualitatively 

evaluated for each identified exposure pathway. 

4.1.1 Identification of Potentially Exposed Human Populations and 

Complete Exposure Pathways

 

Based on the expected use pattern for both types of imidacloprid applications (trunk or soil 

injection), applicators and workers in the program who are mixing and applying the 

insecticide in the field are the most likely subgroup of the human population to be exposed 
Appendix F. Imidacloprid Environmental Risk Assessment F–17

to imidacloprid. Exposure during transportation is not anticipated because the container of 

the concentrated material is sealed. Following label directions including the use of proper 

personal protective equipment (PPE) will minimize exposure to workers. Accidental 

exposure may occur during mixing and applying the formulation. Under an accidental spill 

scenario, workers may be exposed to imidacloprid through dermal contact. However, the 

potential dermal contact exposure is anticipated to be limited because these accidental events 

would be of low frequency and short duration. 

In addition to worker exposure there is the possibility of exposure to residents in areas during 

and after treatment. Generally, exposure to residents during treatment is not expected based 

on the method of application (trunk and soil injection). Exposure to residents whose property 

may be treated can be minimized through proper notification prior to treatment. Therefore, a 

significant exposure pathway is not identified for direct contact to imidacloprid for the 

general public. A significant exposure pathway is not identified for dietary plant 

consumption because treated trees will not have products harvested for human consumption. 

There is the potential for dietary exposure to the public in cases where treated that are used in 

the production of maple syrup; however, this is not a registered use for imidacloprid and this 

type of exposure would not be expected to occur. The soil injection application method may 

result in potential exposure to imidacloprid for children who exhibit pica, the recurrent 

ingestion of unusually high amounts of soil, on the order of 1,000-5,000 mg/day or more 

(USEPA, 2011). In addition, there is the potential for children and adults to be exposed to 

residual imidacloprid in leaf litter. The environmental monitoring data indicates that 

imidacloprid can persist in treated trees for at least 12 months after treatment. Chemical 

uptake occurs throughout the tree with higher residues detected in leaf samples than twig 

samples for most species of trees sampled (USDA APHIS, 2002). 

 

A significant exposure pathway is not idenified for groundwater or surface water media 

under the trunk injection use pattern. Imidacloprid is soluble and has the potential to leach 

from soil to groundwater under the soil injection use pattern. Groundwater sampling 

between 2003 and 2006 in Suffolk County, New York, demonstrated that approximately half 

of the water samples had no detectable levels of imidacloprid. Of those where detections 

occurred, the average concentration was 3.2 parts per billion (ppb) which is below the level 

of concern for human health (USDA APHIS, 2007). Samples with detectable levels of 

imidacloprid do not suggest a contribution from the ALB eradication program because other 

uses of imidacloprid occurred in these areas, and there did not appear to be a significant 

correlation between ALB-related treatment activities and increased residues (USDA APHIS, 

2013). In addition, significant surface runoff into aquatic resources is not expected based on 

the trunk and soil injection use pattern (USDA FS, 2005). 

4.1.2 Exposure Evaluation

 

This section quantitatively evaluates the worker exposures from direct contact pathways 

while applying imidacloprid using trunk and soil injection. 

The proposed use pattern for imidacloprid in the ALB program involves soil injection 

applications using the Merit® 2F, or similar, formulation. The rate of application for 

Merit® 2F is 3-6 ml (0.1-0.2 fl oz) per inch of trunk diameter at breast height with a 
F–18 Appendix F. Imidacloprid Environmental Risk Assessment

maximum use rate of 25.6 fl oz product/acre/year. The formulation is applied in water in 

sufficient quantities to adequately moisten the soil near the base of the tree but not saturate 

the soil which could increase the likelihood of off-site transport. Due to the method of 

application, exposure from drift is not expected. Exposure to applicators will be minimal 

based on the method of application, following the label regarding PPE, and the 

environmental fate of imidacloprid which reduces exposure via inhalation, dermal, and oral 

routes. To quantify the potential exposure to workers making ground applications using 

imidacloprid, estimated doses were derived based on methods developed by the U.S. Forest 

Service, and in particular, those that were developed to assess human health risks to 

imidacloprid (USDA FS, 2005; SERA, 2005). A lower and upper bound estimate of 

exposure was estimated based on assumptions regarding the amount of time and product that 

is applied during a typical work day. 

To quantify the potential exposure to a child from soil ingestion and dermal contact to leaf 

litter, the soil (lower and average concentrations of 0.13 ppm and 1.32 ppm, respectively) and 

leaf litter (average and maximum concentrations of 1.966 ppm and 97.5 ppm, respectively) 

data from the APHIS ALB environmental monitoring program were used. 

4.2 Ecological Exposure Assessment

4.2.1 Aquatic Exposure Assessment

Exposure to aquatic organisms will occur primarily from either runoff related to soil 

applications or from leaf litter that may be deposited into waterbodies. Drift is not expected 

to be an exposure pathway based on the method of application. Estimating residues that 

could occur from leaf deposition into water bodies is difficult to quantify using standardized 

environmental fate models. Aquatic residues from soil injection are based on estimates from 

the U.S. Forest Service that were modeled for various soil types, precipitation levels and 

water bodies (USDA FS 2005). Average estimated aquatic residues in ponds ranged from 

1.49 x 10-7 ppb under low annual rainfall (25 in) to 0.308 ppb under high annual rainfall 

(250 in). Concentrations were less in small streams with average imidacloprid levels of 1.09 

x 10- 8 ppb under low rainfall levels to 0.0829 ppb under high annual rainfall levels. Yearly 

rainfall levels of 25 inches or less resulted in no detectable levels of imidacloprid. This range 

of values is based on the assumption of a one pound per acre application applied to one acre 

of land adjacent to an aquatic habitat. Residues could increase as the amount of acreage 

treated relative to the size of the water body increases; however, as the treatment is moved 

further away from the water body more rainfall would be needed to transport imidacloprid. 

This range of estimated residues was compared to the available effects data for imidacloprid 

in the aquatic risk characterization section of this assessment. 
Appendix F. Imidacloprid Environmental Risk Assessment F–19

4.2.2 Terrestrial Exposure Assessment

Exposure to mammals and birds will be primarily through dietary exposure from consuming 

treated vegetation or invertebrates that consume treated vegetation, or are present in the soil 

where applications may occur. Dermal and inhalation exposure for mammals and birds is not 

expected based on the method of application and environmental fate of imidacloprid. Dietary 

exposure would be greatest for those mammals and birds that feed exclusively on vegetation from 

treated trees. Average and maximum residues for imidacloprid are 1.966 ppm and 97.5 ppm, 

respectively, based on environmental monitoring that has been done by the program. 

Honeybees and other pollinators may be exposed to imidacloprid from nectar and pollen. 

Exposure honeybees to imidacloprid from water is expected to be minor based on the 

methods of application. Field monitoring data from trees treated with imidacloprid during the 

ALB eradication program suggest exposure levels will be low (Johnson, 2012). Pollen samples 

collected from trees treated with imidacloprid through soil injection revealed residues below the 

level of detection (1.0 ppb) to approximately 30.6 ppb with an average concentration of 5.31 

ppb. Greater than 50% of the samples collected from treated trees contained levels below 

detection. Pollen samples collected from trees treated with imidacloprid through trunk injected 

trees were comparably lower, ranging from below detection to 1.5 ppb with an average 

concentration of 0.28 ppb. Greater than two-thirds of the pollen samples had levels below 

detection. The residues collected from this study fall within the low range of imidacloprid 

residues that have been reported in pollen samples from agricultural crops where imidacloprid is 

applied as a seed treatment or a broadcast application (Bonmatin et al., 2015). Pollen residue 

levels are also similar to values reported from hives in various counties in Massachusetts (Lu et 

al, 2015). Nectar samples collected from honeybee hives in the ALB-treated tree study revealed 

no detectable levels of imidacloprid or its metabolites. 

5.0 RISK CHARACTERIZATION

5.1 Human Health

Risks associated with human health are characterized quantitatively for complete exposure 

pathways in this section. Under the PPQ proposed use, imidacloprid use to control ALB 

populations poses minimal risk to human health when applications are made according to label 

directions. 

Estimated exposure doses were based on methods developed for imidacloprid by the U.S. Forest 

Service methods to quantify the potential risk to workers making ground applications of 

imidacloprid (USDA FS, 2005; SERA, 2005). A lower and upper bound estimate of exposure 

was based on assumptions regarding the amount of time and product that is applied during a 

typical work day. The exposure estimate was then compared to the acute (0.14 mg/kg/day) and 

chronic (0.057 mg/kg/day) reference dose to calculate hazard quotient (HQ) values. 

Both acute and chronic HQ values were below one under all scenarios suggesting minimal risk to 

workers making applications of imidacloprid (table 5-1). 

 
F–20 Appendix F. Imidacloprid Environmental Risk Assessment

Table 5-1. Hazard quotient values estimated for workers making imidacloprid applications. 

Lower Estimate Central Estimate Upper Estimate

Hours of Application/Day 6 7 8 

Acres Treated/Day 48 77 144 

Adsorbed Dose Rate 

(mg/kg per lb a.i. /day)

0.00001 0.0002 0.0009 

Adsorbed Dose (mg/kg bw/day) 1.92 x 10-4 6.16 x10-3 5.18 x10-2

Acute Hazard Quotient 0.001 0.04 0.4 

Chronic Hazard Quotient 0.003 0.1 0.9 

 To quantify the HQ from the unusual soil ingestion behavior (pica) and leaf litter ingestion by 

children (age 1-6), soil concentrations were estimated using the application dosage for a soil 

injection scenario. The estimated soil concentrations were compared to the soil concentrations 

collected from the environmental monitoring reports. The imidacloprid concentations from the 

soil monitoring data are higher and were used for the risk estimation. The leaf litter data 

collected from the monitoring reports was also used for estimating risk from those types of 

exposures. HQ values were calculated using the following USEPA soil ingestion risk estimation 

equation for non-carcinogens: 

Acute Hazard Quotient = (Soil/Leaf Litter Concentration x Soil/Leaf Ingestion Rate)/ Body 

Weight or Chronic Hazard Quotient = (Soil/Leaf Litter Concentration x Soil/Leaf Ingestion 

Rate x Exposure Duration x Exposure Frequency x Conversion Factor)/(Averaging Time 

x Body Weight x Reference Dose) (USEPA 2002). 

The calculated acute and chronic HQ values were below one under all scenarios suggesting 

minimal risk to imidacloprid from soil or leaf litter ingestion by children. The upper estimate 

represents the maximum amount of fluid per acre in a single application (table 5-2). 

 Table 5-2. Hazard quotient values estimated for children exposures to soil and leaf litter. 

Lower Estimate Upper Estimate 

Application Dosage for Soil Injection less than 6.0 fl oz per acre in a single application and 2 

pounds of imidacloprid per gallon 

Estimated Soil Concentration 0.0445 mg/kg 0.089 mg/kg

Soil Monitoring Data 0.13 mg/kg 1.32 mg/kg

Leaf Litter Concentration 

Acute Hazard Quotient 

Soil 

Leaf Litter 

1.966 mg/kg 

0.1 

0.02

97.5 mg/kg 

1 

1

Chronic Hazard Quotient 

Soil 

Leaf Litter

0.0008 

0.01

0.008 

0.7

Note: A default soil bulk density of 1.55 g/cm3 for silt loams, and silty clay loams soil (USDA NRCS, 2014)

was used to estimate soil concentration. 
Appendix F. Imidacloprid Environmental Risk Assessment F–21

The risks for the general public from potential exposure to imidacloprid during application are not 

quantified because of minimal potential for exposure and low exposure frequency and duration if 

potential exposure occurs. Dietary consumption of maple syrup from trees treated with 

imidacloprid is not expected because that is not a labelled use and those trees would not receive 

treatment. Consequently, the potential adverse health risks for the general public are expected to 

be minimal. 

The risks to the general public from potential exposure to imidacloprid released to soil and 

transported to surface water or groundwater are not quantified because of the low surface runoff 

potential based on the proposed use pattern. 

5.2 Aquatic and Terrestrial Ecological Risk Characterization

This section of the risk assessment will integrate the effects analysis, regarding toxicological 

studies summarized for terrestrial and aquatic nontarget organisms, with the potential for 

exposure, to quantify the risk to nontarget organisms and domestic animals from program use of 

imidacloprid. Direct risk refers to those risks that could occur from direct exposure to 

imidacloprid while indirect risk refers to impacts that could occur to prey or habitat that 

nontarget organisms rely on for food and shelter. Impacts to individuals and populations are not 

anticipated in cases where the range of imidacloprid toxicity data to imidacloprid does not 

overlap the range of potential exposure values. Further evaluation of the assumptions used in the 

risk characterization is required to refine risk where residues exceed the response data. 
F–22 Appendix F. Imidacloprid Environmental Risk Assessment

5.2.1 Aquatic

The potential risk of imidacloprid applications to aquatic environments is low for tree and 

soil injections. The low risk to these types of habitats is related to the methods of application 

proposed for ALB where trees are injected directly with imidacloprid or the chemical is 

injected into the soil under the dripline of the tree. Drift and runoff are not anticipated to be 

major pathways for off-site transport of imidacloprid to aquatic environments. There is a 

greater chance for runoff from soil injections compared to tree injections; however, the label 

restrictions and estimates of residues when compared to the available toxicity data show the 

risk to be low (Figure 2). The range of effects data discussed in the response section was 

compared to the residue values estimated by the U.S. Forest Service for soil injections in 

various soil types, and water bodies, under various annual rainfall amounts. The comparison 

of the two data sets suggests that acute and chronic risk to aquatic organisms from soil 

application of imidacloprid would be low based on the assumptions used in the exposure 

modeling. 

 

Figure 2. Aquatic risk characterization for imidacloprid soil applications. 

 The greatest potential for residues to reach aquatic habitats is from the leaves of treated trees. 

Acute risk of direct effects to fish and amphibians is very low with an increase in risk to 

aquatic invertebrates from this pathway. Impacts to aquatic invertebrates can result in 

indirect effects to fish and amphibians that rely on invertebrates as a food source. Previous 

research has indicated that lethal effects from this pathway are not expected based on field 

observed levels of imidacloprid in leaf litter deposited in water bodies (Kreutzweiser et al., 

2007; Kreutzweiser et al., 2008a; Kreutzweiser et al., 2009). However some sublethal 

impacts, such as cessation of feeding, have been noted (Kreutzweiser et al., 2009). 

Estimated Acute Aquatic

Sublethal/Chronic

Acute Fish

Sublethal/Chronic

Acute

0

Log Imidacloprid
Appendix F. Imidacloprid Environmental Risk Assessment F–23

5.2.2 Terrestrial Wildlife and Domestic Animals

The direct risk of imidacloprid to mammals and birds is expected to be low based on the 

available toxicity data and the method of application and label requirements that reduce the 

potential for dietary exposure. This would include domestic animals such as dogs and cats 

which would not be expected to forage on treated plants or soil under trees that have been 

treated. Vegetation collected from previous program treatments show average and 

maximum concentrations of 1.966 ppm and 97.5 ppm, respectively. Plausible exposure 

scenarios for wild mammals and birds from either tree injections or soil treatments is difficult 

to quantify; however, the direct risk to nontarget wildlife is expected to be low based on the 

available toxicity data and methods of application (USDA FS, 2005). Not all trees and other 

plant food sources for wild mammals and birds will be treated; therefore, exposure to 

imidacloprid treated vegetation will be reduced since other non-treated plants will be within 

the home range of most mammals and birds. 

A recent review of published studies regarding direct and indirect risk to vertebrates from 

imidacloprid use suggests that environmental exposure levels are below concentrations that 

would result in mortality and most reported sublethal effects (Gibbons et al., 2014). There is 

some uncertainty regarding the lack of risk due to the extrapolation of available toxicity data 

and sublethal impacts from one animal group to another or from one use pattern to another. 

Mineau and Palmer (2013) discuss similar direct and indirect risks regarding impacts to birds; 

however, the emphasis of that work was related to seed treatments using imidacloprid as well 

as other neonicotinoid insecticides. Direct risk to birds and mammals from imidacloprid 

treatments in the ALB program would be expected to be comparatively lower than seed 

treatments. Treated seed that is exposed on the ground during planting operations in 

agricultural settings would be more attractive to seed eating birds and mammals resulting in 

greater exposure. In addition, imidacloprid use would occur within a field with little to no 

alternative imidacloprid residue-free seed sources available for foraging birds. This would not 

be the case with imidacloprid treatments from tree or soil injections as part of the ALB 

program. Select trees are treated within a given area based on site-specific needs, and other 

non-treated trees and plants producing seed would be available for foraging by birds and 

mammals. However, there is the possibility that imidacloprid treated trees could result in 

measurable residues in tree seeds. This risk is expected to be low, especially for maples, that 

are not preferred food sources for most seed eating birds and mammals compared to other plant 

species (Rodewald and Abrams, 2002). In addition, maple seeds degrade more quickly than 

hard mast seeds, such as acorns, and would not be available for foraging. Terrestrial 

vertebrates that eat insects and other terrestrial invertebrate species may also be exposed to 

imidacloprid from foraging on species that have consumed plant material from treated trees. The 

risks to terrestrial vertebrates from this type of exposure is expected to be low due to several 

factors. Exposure to most insectivorous vertebrates would be low because there would be 

unexposed invertebrate because the ALB program does not treat all plant material in an area. In 

cases of soil injections, there would also be the potential for soil inhabiting invertebrates that 

occur under the tree dripline to be exposed; however, these would be localized to areas within the 

dripline of an individual tree and would not be area wide. In addition, most vertebrates that feed 

on invertebrates would preferentially consume live prey compared to those that had received 

lethal doses of imidacloprid. 
F–24 Appendix F. Imidacloprid Environmental Risk Assessment

Indirect risk to terrestrial vertebrates through the loss of prey items is also not anticipated to be a 

significant risk because only invertebrates that consume treated material would be potentially 

impacted by imidacloprid. Thus, many invertebrates with no imidacloprid residues would occur 

in an area to provide a food source to insectivorous terrestrial vertebrates. Also, invertebrates 

that occupy treated trees but do not feed on the leaves and twigs would not be impacted by 

imidacloprid. This would also be the case for soil inhabiting invertebrates that may occur under 

the dripline of trees that are treated with imidacloprid using soil injection. Falcone and DeWald 

(2010) evaluated the impacts of imidacloprid broadcast treatments on invertebrate prey items for 

several neotropical bird species in areas where the invasive balsam wooly adelgid occurs. 

Hemipteran and lepidopteran abundance declined in treated areas versus control areas, but 

species richness and composition was similar between treated and nontreated blocks. Total bird 

abundance for the species evaluated did not differ between treated and untreated areas, 

suggesting that these species were able to forage on other arthropods in treated blocks. These 

types of impacts would not be anticipated in the ALB program because imidacloprid is not 

being applied using broadcast applications. 

5.2.3 Terrestrial Invertebrates and Plants

Risks to terrestrial invertebrates will vary based on the method of application. Soil inhabiting 

invertebrates such as earthworms will be at greater risk from imidacloprid soil applications 

compared to tree injections. Imidacloprid residue data that has been collected as part of the 

ALB eradication program from various locations report an average soil concentration of 1.32 

ppm with a maximum value of 43 ppm. Soil injection sites will have elevated imidaclprid levels 

compared to the area under the tree canopy drip line. The reported average soil imidacloprid 

values are within the range of lethal (LC50 > 1.5 ppm) and sublethal (> 0.1 ppm) effects data 

threshold concentrations, suggesting risk to earthworms. The risk to earthworms in the field will 

be affected by site-specific conditions that could impact exposure concentration and duration. 

Any impacts are expected to be restricted to areas where imidacloprid has been injected into the 

soil under the dripline of trees where applications would occur. 

Terrestrial invertebrates that are sensitive to imidacloprid and consume plant material from 

treated trees are expected to be impacted; however, based on the method of application and the 

mode of action for imidacloprid, these risks will be reduced when compared to making 

broadcast applications. Risks to pollinators are expected to be low based on the available 

residue data that has been collected for pollen from trees treated with imidacloprid either 

through trunk or soil injection and the lack of detectable residues in nectar (Johnson, 2012). 

Mean imidacloprid pollen levels were 1.5 and 5.31 ppb in trunk and soil injected trees, 

respectively. These residues are below a majority of the sublethal effects data for similar types 

of exposure. In addition, more than half of the pollen samples in both treatment methods 

contained no detectable levels of imidacloprid and none of the nectar samples contained 

detectable levels of imidacloprid. Pollen data collected from bees in the same study suggest that 

other sources of pollen are being utilized, further reducing exposure to foraging bees, as well as 

other bee life stages. Pollen traps used to collect pollen from foraging bees at the entrance to 

the hives demonstrated that the percentage of pollen from red maples, a preferred ALB-host 

tree, ranged from 16 to approximately 63 percent. Pollen sources from other plants that are not 

treated with imidacloprid will further reduce exposure to foraging bees as well as individuals 

within the hive. Exposure and risk to honeybees that forage on imidacloprid treated trees would 
Appendix F. Imidacloprid Environmental Risk Assessment F–25

be low due to the low residues that have been reported and the frequency of imidacloprid levels 

below detection observed in nectar and pollen. In addition there would also be pollen sources 

from plants that have not been treated with imidacloprid that honey bees would use. 

Direct risk to plants is expected to be low based on the method of application to selective trees, 

and the lack of direct toxicity to terrestrial plants. Indirect risks to terrestrial plants from 

potential impacts to pollinators is also expected to be low based on the available data regarding 

imidacloprid pollen and nectar levels in imidacloprid treated trees. Invertebrates that may feed 

on treated trees, and are considered pollinators, may be impacted if they are sensitive to the 

effects of imidacloprid. However, impacts to the pollination of terrestrial plants would be minor 

because most plants have various pollinators, many of which would not be exposed to any 

imidacloprid residues because they would not consume leaves or twigs from treated trees. 

6.0 UNCERTAINTIES AND CUMULATIVE IMPACTS

The uncertainties associated with this risk evaluation arise primarily from lack of information 

about the effects of imidacloprid, its formulations, metabolites, and potential mixtures to nontarget organisms that can occur in the environment. These uncertainties are not unique to this 

assessment but are consistent with uncertainties in human health and ecological risk assessments 

with any environmental stressor. In addition, there is uncertainty in where an ALB infestation 

may occur in the United States and the extent of imidacloprid use in a given infestation because 

its use is based on site-specific factors. 

Another area of uncertainty is the potential for cumulative impacts to human health and the 

environment from the proposed use of imidacloprid in the ALB program. Areas where 

cumulative impacts could occur are: 1) repeated worker and environmental exposures to 

imidacloprid from program activities in conjunction with other crop use sources; 2) co-exposure 

to other chemicals with a similar mode of action; and 3) exposures to other chemicals in mixtures 

and how that may affect the toxicity of imidacloprid. 

Cumulative impacts may occur from imidacloprid use from other APHIS programs and in 

relation to other chemicals that have a similar mode of action, as well as others that have a 

different mode of action, that can result in synergism, potentiation, additivity or antagonistic 

effects. Other neonicotinoid insecticides include thiamethoxam, clothianidin, dinotefuran, 

nitenpyram, acetamiprid, and thiacloprid (Goulson, 2013). Temporal variability in the 

occurrence of multiple stressors, as well as their effects, are not well understood. As an 

example, available water quality monitoring data in the United States indicate the presence of 

multiple natural and anthropogenic contaminants. Sources for these chemicals can occur from 

point and non-point sources, and the relative contribution from each is dependent on land use in 

a given watershed. Based on the most recent United States Geological Survey National Water 

Quality Assessment (USGS–NAWQA) data for pesticides, frequency of occurrence for two or 

more pesticides in surface water exceeds 80% nationally (Gilliom et al., 2006). When 

considering other organics and trace metals, the combination of mixtures can become extremely 

large, especially when spatial and temporal variability in mixtures that can occur in a given 

watershed are considered. The seasonal variability in mixtures of pesticides and other 

contaminants has been well documented nationally in urban and agricultural areas (Ryberg 

et al., 2010; Gilliom et al., 2006; Stone et al., 2014). An analysis of all detections from 
F–26 Appendix F. Imidacloprid Environmental Risk Assessment

agricultural streams indicated more than 6,000 unique mixtures of 5 pesticides (Gilliom et al., 

2006). As would be expected, based on the large variability in mixtures, the ecological and 

human health response data for these types of exposure scenarios is very limited for all organic 

and inorganic chemicals including those proposed in the program. 

From a human health perspective, the ALB program use of imidacloprid is expected to have 

negligible cumulative impacts. The proposed trunk and soil injection use of imidacloprid in the 

ALB progam is unlikely to enter the food chain or drinking water and be available to 

significantly add to imidacloprid exposure in the human population, or occur with other 

naturally occurring or synthetic compounds that could result in acute or long-term impacts. Use 

of imidacloprid and other neonicotinoids on animals is another exposure source. Repeated 

chronic exposure to imidacloprid from pet medicine may pose possible health risks to 

veterinarians, veterinary technologists, dog caretakers, and owners (Craig et al., 2005). 

However, due to the low probability of imidacloprid exposure to the human population from 

applications as part of the ALB program, the potential for significant cumulative impacts would 

be minor. 

 7.0 REFERENCES

Agatz, A., Ashauer, R. and C.D. Brown. 2014. Imidacloprid perturbs feeding of Gammarus 

pulex at environmentally relevant concentrations. Environmental Toxicology and Chemistry 

33(3):648–653. 

Agatz, A., Cole, T.A., Preuss, T.G., Zimmer, E. and C.D. Brown. 2013. Feeding inhibition 

explains effects of imidacloprid on the growth, maturation, reproduction, and survival of 

Daphnia magna. Environ. Sci. Technol. 47:2909−2917. 

Alexander, A.C., Culp, J.M., Liber, K. and A.J. Cessna. 2007. Effects of insecticide exposure on 

feeding inhibition in mayflies and oligochaetes. Env. Tox Chem. 26(8): 1726-1732. 

Badgujar, P.C., Jain, S.K., Singh, A., Punia J.S., Gupta, R.P., and G.A.Chandratre. 2013. 

Immunotoxic effects of imidacloprid following 28 days of oral exposure in BALB/c mice. 

Environ. Toxicol. Pharm. 35:408-418. 

Bayer (Bayer Environmental Science). 2004. Merit® 2F insecticide specimen label 

http://www.backedbybayer.com/search?utf8=%E2%9C%93&search=Merit%C2%AE+2F+ last 

accessed 3/28/2014. 

Bayer. 2008. Material safety data sheet (MSDS Number: 102000015064 Version 2.1 

Revision Date: 05/19/2008) 

http://www.backedbybayer.com/system/product/msds_en/52/Merit_2F_Insecticide_19May2008. 

pdf, last accessed 3/28/2014. 

Bhattacharya, A. and S.K. Sahu. 2013. Eco-toxicity of neonicotinoid: a case study on the 

impact of imidacloprid using Drawida willsi earthworm as bioindicator. Int. J. Biosci. 3(5): 78- 

86. 
Appendix F. Imidacloprid Environmental Risk Assessment F–27

Blacquière, T., Smagghe, G., vanGestel, C.A.M., and V. Mommaerts. 2012. Neonicotinoids in 

bees: a review on concentrations, side-effects and risk assessment. Ecotoxicology 21:973–992. 

Bortolloti, L., Montanari, R., Marcelino, J., Medrzycki, P., Maini, S., and C. Porrini, 2003. 

Effect of sub-lethal imidacloprid doses on the homing rate and foraging activity of honey bees. 

Bull. Insectol. 56: 63–67. 

Bonmatin, J.M., Giorio, C., Girolami, V., Goulson, D., Kreutzweiser, D.P., Krupke, C., Liess, 

M., Long, E., Marzaro, M., Mithcell, E.A.D., Noome, D.A., Simon-Delso, N., and A. 

Tapparo. 2015. Environmental fate and exposure; neonicotinoids and fipronil. Environ. Sci. 

Pollut. Res. 22: 35-67 

CA DPR (California Department of Pesticide Regulation). 2006a. Environmental fate of 

imidacloprid. revised by Fossen, M., Environmental Monitoring, Department of Pesticide 

Regulation. 16 pp. April 2006. Available at: 

http://www.cdpr.ca.gov/docs/emon/pubs/fatememo/Imidclprdfate2.pdf, last accessed 3/25/2014. 

CA DPR. 2006b. Imidacloprid risk characterization document dietary and drinking water 

exposure, Health Assessment Section, Medical Toxicology Branch, Department of Pesticide 

Regulation. 195 pp.

Canadian Council of Ministers of the Environment (CCME). 2007 Canadian water quality 

guidelines: imidacloprid scientific supporting document, available at: 

http://www.ccme.ca/files/Resources/supporting_scientific_documents/imidacloprid_ssd_1388.pd 

f, last accessed 1/29/2015. 

Costa, C., Silvari, V., Melchini, A., Catania, S., Heffron, J.J., Trovato, A., and R.D. Pasquale. 

2009. Genotoxicity of imidacloprid in relation to metabolic activation and composition of the 

commercial product. Mutation Research 672: 40-44. 

Craig, M.S., Gupta, R.C., Candery, T.D., and D.A. Britton. 2005. Human exposure to 

imidacloprid from dogs treated with Advantage. Toxicol. Mech. Methods 15: 287-291. 

Cresswell, J.E. 2010. A meta-analysis of experiments testing the effects of a neonicotinoid 

insecticide (imidacloprid) on honey bees. Ecotoxicology 20:149-157. 

Fairbrother, A., Purdy, J., Anderson, T., and R. Fell. 2014. Risks of neonicotinoid insecticides 

to honeybees. Env. Toxicol. Chem. 33(4): 719-731. 

Falcone, J.F. and L.E. DeWald. 2010. Comparisons of arthropod and avian assemblages in 

insecticide-treated and untreated eastern hemlock (Tsuga canadensis [L.] Carr) stands in Great 

Smoky Mountains National Park, USA. Forest Ecol. Mgt. 260:856–863. 

Feng, S., Kong, Z., Wang, X., Zhao, L., and P. Peng. 2004. Acute toxicity and genotoxicity of 

two novel pesticides on amphibian, Rana n. hallowell. Chemosphere 56:457–463. 
F–28 Appendix F. Imidacloprid Environmental Risk Assessment

Feltham, H., Park, K., and D. Goulson. 2014. Field realistic doses of pesticide imidacloprid 

reduce bumblebee pollen foraging efficiency. Ecotoxicology 23:317–323. 

Fischer, J., Muller, T., Spatz, A.K., Greggers U., and B. Grunewald. 2014. Neonicotinoids 

interfere with specific components of avigation in honeybees. PLoS ONE 9(3): e91364. 

doi:10.1371/journal.pone.0091364 

Gawade, L., Dadarkar, S.S., Husain, R., and M. Gatne. 2013. A detailed study of developmental 

immunotoxicity of imidacloprid in wistar rats. Food Chem. Toxicol. 51:61-70 

Gervais, J. A., Luukinen, B., Buhl, K., and D. Stone. 2010. Imidacloprid technical fact sheet; 

National Pesticide Information Center, Oregon State University Extension Services. 

http://npic.orst.edu/factsheets/imidacloprid.html last accessed 3/25/2014 

Gibbons, D., Morrissey, C., and P. Mineau. 2014. A review of the direct and indirect effects of 

neonicotinoids and fipronil on vertebrate wildlife. Environ. Sci. Pollut. Res. 

Gilliom, R.J. and P.A. Hamilton. 2006. Pesticides in the Nation’s streams and ground water, 

1992–2001, The Quality of Our Nation’s Waters. U.S. Geological Survey Circular 1291,172 p. 

Graebing P. and J.S. Chib 2004. Soil photolysis in a moistureand temperature-controlled 

environment. 2. Insecticides. Journal of Agricultural and Food Chemistry 52(9): 2606–2614.

Goulson, D. 2013. Review: An overview of the environmental risks posed by neonicotinoid 

insecticides http://onlinelibrary.wiley.com/doi/10.1111/1365-2664.12111/full last accessed 

3/28/2014. 

HSDB (Hazardous Substances Databank). 2014. “Imidacloprid”; IN: Toxnet. U.S. Department 

of Health and Human Services, National Institutes of Health, National Library of Medicine. 

http://toxnet.nlm.nih.gov/cgi-bin/sis/search2/f?./temp/~JQa1Yi:1, last accessed 6/20/2014 

Iyyadurai, R., George, I.A, and J.V. Peter, 2010. Imidacloprid poisoning newer insecticide and 

fatal toxicity, J. Med. Toxicol. 6:77–78 

Jemec, A., Tisler, T., Drobne, D., Sepcic, K., Fournier, D., and P. Trebs. 2007. Comparative 

toxicity of imidacloprid, of its commercial liquid formulation and of diazinon to a non-target 

arthropod., the microcrustacean D. magna. Chemosphere 68:1408–1418. 

Johnson, J.D. 2012. The role of pesticides on honey bee health and hive maintenance with an 

emphasis on the neonicotinoid, imidacloprid. Dissertation submitted to University of Maryland. 

170 pp. 

Key, P., Chung, K., Siewicki, T., and M. Fulton. 2007. Toxicity of three pesticides individually 

and in mixture to larval grass shrimp (Palaemonetes pugio). Ecotox. Environ. Safety 68(2):272- 

277. 

Kreutzweiser, D., Thompson, D., and T.A. Scarr. 2009. Imidacloprid in leaves from 
Appendix F. Imidacloprid Environmental Risk Assessment F–29

systemically treated trees may inhibit litter breakdown by non-target invertebrates. Ecotox. 

and Environ. Safety. 72:1053–1057. 

Kreutzweiser, D., Good, K., Chartrand, D., Scarr, T., and D. Thompson. 2008. Are leaves 

that fall from imidacloprid-treated maple trees to control Asian longhorned beetles toxic to 

non-target decomposer organisms? J. Environ. Qual. 37: 639–646. 

Kreutzweiser, D., Good, K., Chartrand, D., Scarr, T., and D. Thompson 2007. Non-target 

effects on aquatic decomposer organisms of imidacloprid as a systemic insecticide to control 

emerald ash borer in riparian trees. Ecotox. and Environ. Safety. 68: 315–325. 

Kumar, A., Verma, A.B., and A. Kumar. 2013. Accidental human poisoning with a 

neonicotinoid insecticide, imidacloprid: A rare case report from rural India with a brief review of 

literature, Egyptian Journal of Forensic Sciences 3:123-126. http://ac.elscdn.com/S2090536X13000427/1-s2.0-S2090536X13000427-main.pdf?_tid=d32e9344-fd61- 

11e3-9c00-00000aacb360&acdnat=1403808415_f61bd64cdb42a8f2fd88bd50b8749ca8, last 

accessed 6/26/2014. 

Liu, W., Zheng, W., Ma, Y., and K.K. Liu. 2006. Sorption and degradation of imidacloprid in 

soil and water, Journal of Environmental Science and Health Part B, 41:623–634. 

Lin P.C., Lin, H.J., Liao, Y.Y., Guo, H.R., and K.T. Chen. 2013. Acute poisoning with 

neonicotinoid insecticides: a case report and literature review. Basic & Clinical Pharmacology & 

Toxicology, 112, 282–286. http://onlinelibrary.wiley.com/doi/10.1111/bcpt.12027/pdf, last 

accessed 6/26/2014. 

Lu, C., Chang, C., Tao, L. and M. Chen. 2015. Distributions of neonicotinoid insecticides in the 

Commonwealth of Massachusetts: a temporal and spatial variation analysis for pollen and honey 

samples. Environ. Chem. http://www.publish.csiro.au/?paper=EN15064. 

Maus, C., Cure, G., and R. Schmuck. 2003. Safety of imidacloprid seed dressings to honey 

bees: a comprehensive overview and compilation of the current state of knowledge. Bull. 

Insectol. 56(1): 51–57. 

Hrouzková, S. and E. Matisova. 2012. Chapter 5 Endocrine disrupting pesticides, Pesticides - 

Advances in Chemical and Botanical Pesticides. http://www.intechopen.com/books/pesticidesadvances-in-chemical-and-botanical-pesticides/endocrine-disrupting-pesticides last accessed 

3/27/2014. 

Mineau, P. and C. Palmer. 2013. The impact of the nation’s mostwidely used insecticides on 

birds. American Bird Conservancy, USA. 97 pp. 

Morandin, L.A. and M.L. Winston. 2003. Effects of novel pesticides on bumble bee 

(Hymenoptera: Apidae) colony health and foraging ability. Environ. Entomol. 32(3): 555– 

563. 

Mohamed, F., Gawarammana, I., Robertson, T.A., Roberts, M.S., Palangasinghe, C., Zawahir, 
F–30 Appendix F. Imidacloprid Environmental Risk Assessment

S., Jayamanne, S., Kandasamy, J., Eddleston, M., Nick A. Buckley, N.A., Dawson, A.H., and 

D.M. Roberts. 2009. Acute human self-poisoning with imidacloprid compound: a neonicotinoid 

insecticide. PLoS ONE 4(4): e5127. doi:10.1371/journal.pone.0005127, 

http://www.plosone.org/article/info%3Adoi%2F10.1371%2Fjournal.pone.0005127, last accessed 

6/20/2014 

Mohany, M., Badr, G., Refaat, I., and El-Feki, M. 2011. Immunological and histological 

effects of exposure to imidacloprid insecticide in male albino rats. African J Pharm Pharmacol 

5(18): 2016-2114. 

Najafi, G., Razi, M., Hoshyar, A., Shahmohamadioo, S., and S. Feyzi. 2010. The effect of 

chronic exposure with imidacloprid insecticide on fertility in mature male ats. International 

J.Fertility Sterility 4(1):9-16. 

NRC (National Research Council). 1983. Risk assessment in the Federal government: 

managing the process. National Academy Press, Washington, DC. 

Nyman, A.M., Hintermeister, A., Schirmeri, K., and R. Ashauer. 2013. The insecticide 

imidacloprid causes mortality of the freshwater amphipod, Gammarus pulex by interfering with 

feeding behavior. PLoS ONE 8(5): e62472. doi:10.1371/journal.pone.0062472. 

Overmyer, J.P., Mason, B.N., and K.L. Armbrust. 2005. Acute toxicity of imidacloprid and 

fipronil to a nontarget aquatic insect, Simulium vittatum Zetterstedt cytospecies IS-7. Bull. 

Environ. Cont. Tox. 74(5):872-879. 

Paul, A., Harrington, L.C., and J. G. Scott. 2006. Evaluation of novel insecticides for control of 

dengue vector Aedes aegypti (Diptera: Culicidae). J. Med. Ent. 43(1):55-60. 

Pestana, J.L.T., Alexander, A.C., Culp, J.M., Baird, D.J., Cessna, A.J., and A.M.V.M. Soares. 

2009. Structural and functional responses of benthic invertebrates to imidacloprid 

in outdoor stream mesocosms. Environ. Poll. 157:2328–2334. 

Pisa, L.W., Amaral-Rogers, V., Belzunces, L.P., Bonmatin, J.M., Downs, C.A., Goulson, D., 

Kreutzweiser, D.P., Krupke, C., Liess, M., McField, M., Morrissey, C.A., Noome, D.A., Settele, 

J., Simon-Delso, N., Stark, J.D., van der Sluijs, J.P., VanDyck, H., and M. Wiemers. 2015. 

Effects of neonicotinoids and fipronil on non-target invertebrates. Environ. Sci. Pollut. Res. 

22:68–102. 

Rodewald, A.D. and M.D. Abrams. 2002. Floristics and avian community structure: 

implications for regional changes in eastern forest composition. For. Sci. 48(2): 267-272. 

Roessink, I., Merga, L.B., Zweers, H.J., and P.J. Vanden Brink. 2013. The neonicotinoid 

imidacloprid shows high chronic toxicity to mayfly nymphs. Env. Toxicol.Chem. 32(5): 1096– 

1100. 

Ruiz de Arcaute, C., Perez-Iglesias, J.M., Nikoloff, N., Natale, G.S., Soloneski, S., and M.L. 
Appendix F. Imidacloprid Environmental Risk Assessment F–31

Larramendy. 2014. Genotoxicity evaluation of the insecticide imidacloprid on circulating blood 

cells of Montevideo tree frog, Hypsiboas pulchellus tadpoles (Anura, Hylidae) by comet and 

micronucleus bioassays. Ecol. Indicators 45:632–639. 

Ryberg, K.R., Vecchia, A.V., Martin, J.D., and R.J. Gilliom. 2010. Trends in pesticide 

concentrations in urban streams in the United States, 1992–2008: U.S. Geological Survey 

Scientific Investigations Report 2010–5139, 101 p. 

Sánchez-Bayo, F. and K. Goka. 2005. Unexpected effects of zinc pyrithione and imidacloprid 

on Japanese medaka fish (Oryzias latipes). Aquatic Tox. 74.4: 285-293 

Sarkar, M., S. Roy, Kole, R., and A. Chowdhury. 2001. Persistence and metabolism of 

imidacloprid in different soils of West Bengal. Pest Manag. Sci. 57: 598—602. 

Schmuck, R. 2004. Effects of a chronic dietary exposure of the honeybee, Apis mellifera

(Hymenoptera: Apidae) to imidacloprid. Arch. Environ. Contam. Toxicol. 47: 471–478. 

Schmuck, R., Schoning, R., Stork, A., and O. Schramel. 2001. Risk posed to honeybees (Apis 

mellifera L., Hymenoptera) by an imidacloprid seed dressing of sunflowers. Pest Manag. Sci. 

57: 225–238.

Scholer, J. and V. Krischik. 2014. Chronic exposure of imidacloprid and clothianidin reduce 

queen survival, foraging, and nectar storing in colonies of Bombus impatiens. PLoS ONE 9(3): 

e91573. doi:10.1371/journal.pone.0091573. 

SERA (Syracuse Environmental Research Associates, Inc.). 2005. Documentation for 

Worksheets Version 4.01 - Human Health and Ecological Risk Assessments. Report dated May 

4, 2005. Available at: www.sera-inc.com. 

Song, M.Y., Stark, J.D., and J.J. Brown. 1997. Comparative toxicity of four insecticides, 

including imidacloprid and tebufenozide, to four aquatic arthropods. Env. Toxicol. Chemistry 

16:(12):2494-2500. 

Song, M.Y. and J.J. Brown. 1998. Osmotic effects as a factor modifying insecticide toxicity on 

Aedes and Artemia. Ecotox. Environ. Safety 41(2):195-202. 

Stadler, T., Gines, D.M., and M. Buteler. 2003. Long-term toxicity assessment of imidacloprid 

to evaluate side effects on honey bees exposed to treated sunflower in Argentina. Bull. Insectol. 

56(1): 77–81. 

Stone, W.W., Gilliom, R.J., and J.D. Martin. 2014, An overview comparing results from two 

decades of monitoring for pesticides in the Nation’s streams and rivers, 1992–2001 and 2002– 

2011: U.S. Geological Survey Scientific Investigations Report 2014–5154, 23 p., 

http://dx.doi.org/10.3133/sir20145154. 

Stoughton, S.J., Liber, K., Culp, J., and A. Cessna. 2008. Acute and chronic toxicity of 

imidacloprid to the aquatic invertebrates Chironomus tentans and Hyalella azteca under 

constant- and pulse-exposure conditions. Arch Environ Contam Toxicol 54: 662–673 
F–32 Appendix F. Imidacloprid Environmental Risk Assessment

Tasei, J.N., Lerin, J. and G. Ripault. 2000. Sublethal effects of imidacloprid on bumblebees, 

Bombus terrestris (Hymenoptera: Apidae), during a laboratory feeding test. Pest Manag. Sci. 

56: 784–788. 

Tasei, J.N., Ripault, G., and E. Rivault. 2001. Hazards of imidacloprid seed coating to Bombus 

terrestris (Hymenoptera: Apidae) when applied to sunflower. J. Econ. Entomol. 94(3): 623–627. 

Tasei, J.N. 2002. Impact of agrochemicals on non-Apis bees. In: Honey bees: estimating the 

environmental impacts of chemicals. Eds. J. Devillers and M.H. Pham-Delegue. Taylor and 

Francis Publishing. pp. 101–131. 

Tomizawa, M. and J.E. Casida. 2003. Selective toxicity of neonicotinoids attributable to 

specificity of insect and mammalian nicotinic receptors. Annu. Rev. Entomol. 48:339-364. 

USDA APHIS (United States Department of Agriculture, Animal and Plant Health Inspection 

Service). 2001. Environmental monitoring report, 2000 - 2001 Asian Longhorned Beetle 

Cooperative Eradication Program prepared by Enviornmental Monitoring Team. 

USDA APHIS. 2002. Asian Longhorned Beetle Cooperative Eradication Program In New York 

and Illinois environmental monitoring report 2002, prepared by Environmental Monitoring 

Team. 

USDA APHIS. 2007. Environmental monitoring report: 2006 Asian Longhorned Beetle 

Cooperative Eradication Program for the active eradication region in Suffolk County, New York. 

36 pp. 

USDA APHIS. 2013. Asian Longhorned Beetle Cooperative Eradication Program in Clermont 

County, Ohio Revised Environmental Assessment, May 2013 

USDA FS (United States Department of Agriculture, Forest Service). 2005. Imidacloprid: 

human health and ecological risk assessment (final report). SERA TR 05-43-24-03a. 283 pp. 

USDA NRCS (United States Department of Agriculture. National Resource Conservation 

Service). 2014. Soil Bulk Density/Moisture/Aeration. 

http://www.nrcs.usda.gov/Internet/FSE_DOCUMENTS/nrcs142p2_053260.pdf, last accessed 

6/26/2014. 

USEPA (United States Environmental Protection Agency). 1989. Risk assessment guidance for 

Superfund volume I, human health evaluation manual (Part A). interim final. Office of 

Emergency and Remedial Response, Washington, DC. EPA/540/1-89/002 December 1989. 

USEPA. 2002. Supplemental guidance for developing soil screening levels for Superfund sites, 

Solid Waste and Emergency Response, Office of Emergency and Remedial Response, 

Washington DC, OSWER 9355.4-24 

USEPA. 2005. Imidacloprid; pesticide tolerances for emergency exemptions. Fed Regist. 
Appendix F. Imidacloprid Environmental Risk Assessment F–33

70(16): 3634-3642. January 26, 2005. Available At: http://www.epa.gov/fedrgstr/EPAPEST/2005/January/Day-26/p1438.htm 

USEPA. 2008. OPP Ecotoxicity Database. Available at: 

http://www.ipmcenters.org/Ecotox/index.cfm.

USEPA. 2010. Imidacloprid: revised human-health risk assessment for proposed Section 3 seed 

treatment uses on bulb vegetables (Crop Group 3); cereal grains (Crop Group 15); root and tuber 

vegetables, except sugar beet (Crop Subgroup IB); tuberous and corm vegetables (Crop 

Subgroup I C); leafy vegetables, except Brassica (Crop Subgroup 4A); Brassica vegetables 

(Crop Group 5); fruiting vegetables (Crop Group 8); cucurbit vegetables (Crop Group 9), and 

residential crack and crevice and bed-bug uses. Petition Nos: 8F7414, 8F7415, Section 3 

Registration, Memorandum from Kramer et al. of Risk Assessment Branch, Health Effects 

Division to Kable Davis/Venus Eagle of Registration Division, dated March 16, 2010, available 

at: http://www.epa.gov/pesticides/chem_search/hhbp/R181434.pdf, last accessed 2/10/2015. 

USEPA. 2011. Exposure factor handbook: 2011 Edition. Office of Research and Development, 

Washington, DC, National Center for Environmental Assessment, EPA/600/R-090/052F 

USEPA. 2013. Chemicals evaluated for carcinogenic potential, Office of Pesticide Programs, 

http://npic.orst.edu/chemicals_evaluated.pdf, last accessed June 20, 2014 

USEPA. 2014a. Overview of Risk Assessment in the Pesticide Program – Human Health Risk 

Assessment, Office of Pesticide Programs 

(http://www.epa.gov/pesticides/about/overview_risk_assess.htm) last accessed 3/21/2014. 

USEPA. 2014b. Endocrine Disruptor Screening Program, Office of Chemical Safety and 

Pollution Prevention, Office of Science Coordination and Policy (http://www.epa.gov/endo/), last 

accessed 3/27/14. 

USEPA. 2014c. List of contaminants and their maximum contaminant levels, Water: Drinking 

Water Contaminants. http://water.epa.gov/drink/contaminants/, last accessed 6/24/2014. 

Van der Sluijs, J.P., Simon-Delso, N., Goulson, D., Maxim, L., Bonmatin, J.M, and L.P 

Belzunces. 2013. Neonicotinoids, bee disorders and the sustainability of pollinator 

services. Current Opinion in Environmental Sustainability. 5:293–305. 

Westwood, F., Bean, K.M., Dewar, A.M., Bromilow, R.H., and K. Chamberlain. 1998. 

Movement and persistence of (14C) imidacloprid in sugar-beet plants following application to 

pelleted sugar-beet seed. Pesticide Science. 52(2):97-103. 
F–34 Appendix F. Imidacloprid Environmental Risk Assessment

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Appendix G. Response to Comments G–1 

Appendix G. Response to Comments on the Asian 

Longhorned Beetle Cooperative 

Eradication Program, Draft Environmental 

Impact Statement, March 2015 

This appendix provides a response to comments received on the March 2015 draft EIS APHIS 

published in the Federal Register on March 16, 2015 (Federal Register, Volume 80, Number 49). 

Several comments point out text in the EIS that needs clarification or editorial correction. 

Corrections were made within the document, where appropriate, while comments were either 

addressed directly in the document with additional text and/or in this appendix. No comments 

trigger significant changes to the alternatives, environmental consequences, or cumulative impacts 

in the final EIS. 

Comment 1 

Several commenters found that the draft EIS was not descriptive enough on the restoration 

practices following ALB host removal. One commenter observed that the Program does not 

necessarily replant shade trees according to the Program’s practice of restoring the area after tree 

removal. Another commenter mentions that replanting efforts during full host removal are lacking. 

“Many of the options in C. Decision Framework, mentions restoration of removed trees. Yet the 

restoration process is not mentioned in any detail in this document. That information could be 

useful to help myself and the public understand and appreciate the efforts the program is taking to 

help restore areas as this information is lacking in the document.” 

After tree removal, the Program stabilizes the site to reduce erosion by planting a groundcover 

(e.g., grass) and may grade the site. Planting a groundcover also reduces the opportunity for 

invasive weeds to establish. APHIS follows best management practices (BMP) for erosion control 

to the extent possible, considering weather conditions. For example, in Ohio, the ALB program 

adheres to the BMP for Erosion Control for Logging Practices in Ohio (Bulletin 916). In addition 

to this, any impacts will be mitigated to the extent that USDA, cooperators, and stakeholders assist 

with replanting trees or managing impacted areas.

Comment 2 

One commentor questioned whether the Program’s assessment on the impact of tree removal to the 

value of residential properties and infrastructure, in particular, the breaking apart of roads since 

water is not absorbed by the many trees taken from the area, is properly addressed in the draft EIS. 

In the draft and this final EIS, APHIS considers the economic and environmental values trees 

contribute to residential properties and communities. APHIS recognizes the impact from tree 

removal may be greater for those properties which have a tree cover dominated by ALB-host trees. 

The draft and final EIS discuss the important role trees have in water quantity and quality as well 

as other important characteristics such as shading and erosion control. In general, damage to road 

surfaces may occur in areas with high traffic volume, frequent use by heavy vehicles, and weather 

events such as freeze-thaw cycles.Tree roots can also distress sidewalks and roads causing cracks, 
G–2 Appendix G. Response to Comments 

vaults, and faults. The quality of the road construction materials and the road’s design will affect 

the road’s rate of decline (Charlier et al., 2009). The removal of trees in the ALB eradication 

program may contribute to road deterioration in some areas through the use of heavy equipment to 

remove the trees; however, APHIS expects the contribution to be minimal given the existing traffic 

volumes on roadways and the short duration heavy equipment used in the Program will travel on 

roads in the impacted area. Roads consist of several layers of different porosity which aid in water 

permeability (Charlier et al., 2009). Road design for porosity is usually independent of the water 

uptake from trees. Water accumulation can cause surface deterioration (Randup et al, 2001 cites 

Army Corps of Engineering; Charlier et al., 2009). The removal of trees in the ALB Program may 

cause an initial increase in water availability in the area through the reduction of plant biomass 

available to uptake water (Randrup et al., 2001). After tree removal, the ALB Eradication Program 

plants a groundcover to reduce erosion and reduce the growth of invasive plant species through 

competition. The ALB Eradication Program encourages homeowners and municipalities to plant 

non-ALB-host plants. In the final EIS, APHIS discusses the impact to road infrastructure from ALB 

program activities. 

Comment 3 

One commenter points out that not all research is complete regarding the use of chemical treatment 

to manage ALB. 

In both the draft and final EIS, APHIS addresses impacts and uncertainties associated with the 

various ALB management strategies, including the use of chemical treatments in the Program. 

APHIS adopts adaptive management which allows the integration of management strategies with 

equivalent or better effectiveness and equivalent or less impact to human health and ecological 

resources. APHIS has conducted research regarding the efficacy of other insecticides for the 

treatment of ALB. To date, imidacloprid has proven the most effective method for prophylactic 

treatment of ALB host trees. APHIS will continue to evaluate other treatment methods, including 

insecticide treatments, and will incorporate new methods as part of its adaptive management 

strategy. 

Comment 4 

Several commenters mention the importance of wood packaging material with imports as a 

continued pathway for ALB introduction into the United States. One commentor, although 

supporting the alternative option #5 – an integrated approach, thinks the eradication stategy will 

fail unless regulatory agencies take a more proactive approach to prevent new populations of ALB 

from entering North America. The commenter identifies wood packaging material (pallets and 

crates), despite being heat treated or fumigated, as a viable pathway for ALB. One commenter 

suggests “The brief discussion of the potential threat of ALB spread associated with wood 

packaging used in domestic U.S. shipments should note that APHIS has no plans to regulate this 

pathway other than through quarantine restrictions which the agency concedes to be leaky.” 

USDA works closely with the Department of Homeland Security to ensure that measures are in 

place to reduce risk of introduction and establishment of pests from imported commodities. 

International trade has increased markedly in recent history. As a result, risk of invasive pests has 

increased. Inspections at ports are subject to available resources and require prioritization of 
Appendix G. Response to Comments G–3 

items for inspection. ALB likely entered the United States from Asia in solid wood packing 

materials used to transport trade goods. Several of the ALB infestations in the United States (e.g., 

Ohio around 2004) occurred before the adoption of the international regulations for solid wood 

packing materials in North America in 2006 (International Plant Protection Convention, 

International Standards for Phytosanitary Measures 15). These regulations reduce the risk of 

further ALB introductions into the United States. 

Both State and Federal law establish regulated areas around ALB infestations. The Federal 

regulation can be found in the Code of Federal Regulations (CFR), specifically in 7 CFR 

301.51. Requirements for movement of regulated articles are outlined in those regulations. When 

an inspector has probable cause to believe a person is moving a regulated article interstate, the 

inspector is authorized to stop the person to determine whether a regulated article is present and 

to inspect it. The inspector may seize, quarantine, treat, destroy, dispose, or subject to other 

remedial measures infested articles. 

Comment 5 

One commenter summarizes several concerns regarding erosion and sediment control, 

consideration of fugitive dust, and minimizing emissions during periods of high ozone. 

The ALB Eradication Program minimizes soil erosion and sediment by grading tree removal sites 

and planting a groundcover. The Program also follows best management practices for tree 

removal, where appropriate. The EIS considers erosion in the draft and final EIS. APHIS discusses 

particulate matter, which includes fugitive dust, in the Affected Environment chapter. In the final 

EIS, APHIS adds the term “fugitive dust” to the air quality and environmental consequences 

sections. The EIS considers air pollutants, including ozone; however, the EIS does not specifically 

mention minimizing emissions during periods of high ozone. APHIS will continue to work with 

state agencies, when appropriate, regarding any required permits related to air quality to ensure 

program activities do not result in impacts to air quality. APHIS recognizes that emissions from 

program activities may occur during times of high ozone however the related emissions are 

expected to be minor relative to other activities that may be occurring. 

Comment 6 

One commenter informs APHIS that imidacloprid and other neonicotinoid insecticides are 

undergoing re-evaluation in the EPA registration review program which may lead to changes in the 

use of this class of insecticides. There was also a comment regarding the potential impacts of 

imidacloprid to pollinators and uncertainty about impacts from tree applications. 

APHIS is aware of the re-evaluation by EPA of the neonicotinoid insecticide class as it relates to 

potential pollinator impacts. APHIS follows all current label instructions and restrictions for 

pesticides and complies with the Fungicide, Insecticide, and Rodenticide Act (FIFRA). Proposed 

changes to labels as a result of this re-evaluation would be implemented by the ALB Eradication 

Program. APHIS recognizes that there may be an increased risk to pollinators from the use of 

imidacloprid and other neonicitonoid insecticides. Information regarding impacts to pollinators is 

summarized in the draft and final EIS with more specific information in the Appendix F which 
G–4 Appendix G. Response to Comments 

discusses the human health an ecological risk of imidacloprid use in the ALB Eradicaiton 

Program. As summarized in the appendix, APHIS has funded studies to evaluate the potential risk 

of imidacloprid from ALB treatments to address some of the uncertainty regarding the potential for 

exposure and risk to pollinators. 

Comment 7 

Several commenters provide statements in support of APHIS recommendation of the preferred 

alternative. 

Comment noted. 

Comment 8 

One commenter found the terminology misleading on how imidacloprid treatment protects trees. 

Specifically, the commenter indicates the use of the “prophylactic” is incorrect as well as the 

statement that imidacloprid protects trees. The commenter suggests replacing the words “protecting 

trees” with “killing ALB”. 

The term prophylactic has been used in the scientific literature to describe how imidacloprid can 

be used to control ALB. Its use in this example is to show that imidacloprid use is not a curative 

treatment against ALB (ie. controlling ALB once a tree has become infested) but can be used to 

eliminate ALB that feed on insecticide-treated trees prior to becoming infested. 

Comment 9 

Several comments pertain to the Program’s outreach activities and recommend APHIS get more 

people volunteering to help with ALB Program activities. “Volunteering is a great 

solution/opportunity for agencies and the public. My familiarity of the program allows for 

compliance training and a hotline to report a beetle or damage in a tree. Do allow people to help. 

Do not brush the public away, and make it a priority.” 

APHIS recognizes the importance of public involvement in invasive species activities. In fact, the 

first report of ALB in the United States came from a homeowner who noticed the beetle in their 

backyard. APHIS recently issued a press release encouraging homeowners to look for ALB signs 

since August is a time of peak emergence making it easier to observe potential infestations. APHIS 

encourages public involvement in the ALB eradication program and other pest eradication 

programs but acknowledges the importance of public safety in eradication activities such as tree 

removal. APHIS will continue to work with the public and other stakeholders regarding ALB 

Eradication Program activities. 

Comment 10 

Several comments pertain to the need for more research and data on the biology and management 

of ALB. Another comment indicates that referenced research is old, that more research should be 
Appendix G. Response to Comments G–5 

done and newer studies should be used. More information on current/ongoing research should be 

included. 

The USDA and other groups fund research in ALB, including research in pest detection and 

management. APHIS incorporates the latest science into their programs, including the ALB 

Eradication Program. APHIS is also using adaptive management which allows the integration of 

management strategies with equivalent or better effectiveness and equivalent or less impact to 

human health and ecological resources. 

Comment 11 

One commenter suggest a discussion on the threat to urban and rural forests next to existing 

infestations, specifically to Kentucky which is close to the ALB quarantine area in Clearmont 

County, Ohio. 

In the draft and final EIS, APHIS summarizes factors that may make an area at risk of infestation, 

such as the availability of host trees and proximity to an existing ALB infestation. This is a 

programmatic EIS and site specific factors will be addressed in site specific EAs. 

Comment 12 

One commenter suggests the EIS provide additional information on how the shift in trees species 

in some areas from ALB host trees to non-host trees as a result of climate change does not indicate 

a more healthy forest since other tree types are vulnerable to pests. 

In the draft and final EIS, APHIS provides this information.

Comment 13 

One commenter suggests APHIS strengthen the section on firewood as a vector for human-assisted 

spread of ALB. 

APHIS identifies firewood as a pathway for the domestic movement of some invasive pests, 

including the ALB. In the draft and final EIS, APHIS mentions firewood as a pathway and 

describes the quarantine restrictions for firewood. 

Comment 14 

One commenter suggests adding greater specificity about the threat to individual cities, as some 

cities have both a high proportion of trees in the urban forest that are vulnerable to ALB and a high 

risk of introduction of the insect due to receiving large volumes of imports from Asia. 

APHIS pest surveillance activities (not limited to ALB) consider various pathways of pest entry 

into the United States and the likelihood of establishment. The programmatic EIS does not evaluate 

the risks of pest introduction from offshore sources or likelihood of establishment from these 

introductions but evaluates the impact of APHIS actions in the eradication of ALB. 
G–6 Appendix G. Response to Comments 

Comment 15 

One commenter suggests expanding on the cost burden on municipal governments to remove and 

replace trees. 

APHIS agrees that munipal governments may experience cost burdens associated with the removal 

and replacement of trees. In this final EIS, APHIS expands on the potential direct and indirect 

economic impacts to municipalities and landowners from the removal of trees on their property. 

Comment 16 

“Enhanced recognition that many of the forests threatened by ALB spread are already affected by 

other pests – so the loss of ALB hosts would exacerbate already serious forest health problems. 

[...] While the ALB eradication program does not deal with these threats, the DEIS discussion of 

cumulative effects should address their implications. Such a discussion is especially important as 

part of the description of the importance of maintaining forests in riparian areas. There might also 

be generalizable impacts from additive loss of soft mast. The list of sources includes articles on 

this cumulative pest impacts by Dr. Canham and others (including APHIS staff), but the text does 

not appear to include the full breadth of these studies.” 

The draft and final EIS discusses other forest health problems and the cumulative effects of ALB 

and the ALB eradication program within the context of these problems. The EIS also discusses 

riparian areas as well as impacts to wildlife food resources through the loss of ALB host trees. 

Comment 17 

“Can the DEIS explore more fully the implications of data and correlations in Clermont County, 

Ohio that appear to link ALB establishments more closely to humidity levels than simply to 

temperatures? Are there regions of the country that might be either too hot during the summer 

(sustained temperatures at or above 30o

C/86oF) or too dry to support ALB? Of course, we are not 

advocating that APHIS or others should relax their vigilance or abandon the eradication program.” 

Predictive habitat distribution models incorporate climate and other environmental factors to 

understand the potential distribution of an organism (Guisan and Zimmermann, 2000). Models are 

a simulation of what might be but there are uncertainties with every model. In the draft and final 

EIS, APHIS summarizes from the scientific literature a few ALB distribution models, all of which 

indicate that ALB is likely to survive anywhere in the conterminous United States where host trees 

occur. As mentioned in the draft and final EIS, it is possible that the host range for ALB may 

change. Also, the temperature scale used in the models does not take into account microclimate 

(temperature and moisture) variations which can affect an organisms survival and population 

growth. The conditions inside the tree will differ from that outside the tree and will likely provide 

protection to ALB larvae and pupae during extreme climatic events such as high ambient 

temperatures or drought. APHIS expects predictive distributions, including those for ALB, to 

change as more information becomes available on the interaction between environmental factors 

and a pest or pathogen’s biology. 

 
Appendix G. Response to Comments G–7 

Comment 18 

A couple commenters suggest utilizing ALB movement patterns to better allocate ALB eradication 

program resources. Reporting and tracking the number of beetles captured may be helpful in 

allocating resources and focusing project efforts. “Use catch and release studies as a population 

estimate or other analytical methods to predict ALB presence. We need to see if we can refine 

ALB movement patterns to help put resources in the right place.” 

The ALB Eradication program has modified its strategies in response to new scientific information 

as well as experience in the program’s operation. The ALB survey design takes into account ALB 

movement patterns. However, estimating the population density may prove difficult because we 

currently do not have a lure to trap adult ALB. Detection of ALB in the initial few years of an 

outbreak is difficult because the beetles have low fecundity, infested trees may not display 

symptoms within the first few years of infestation, and the specific signs (eggs sites, adult exit 

holes, frass, etc.) can be difficult to visually detect. 

Comment 19 

“Every regulated area should allow the public to use the dump site: will provide many benefits 

including an eradication program that is more informed about what’s going on with the host 

material in and around the regulated area. It is a great carrot especially since states don’t follow 

through on enforcement fines.” 

APHIS agrees that the proper movement and disposal of regulated host material is important in 

preventing the introduction and establishment of ALB outside of the quarantine area. 

Comment 20 

One commenter suggests that “ALB could work with existing private and government weed 

management entities to better mitigate invasive weed establishment, as well as more carefully 

following their states BMPs which should provide guidelines for controlling invasive plants.” 

Another comment related to invasive weeds growing in tree removal sites states “If the ALB 

program is going to spend millions of dollars to prevent the spread of one invasive species, we 

have to account for the additional cost of other invasive species it is helping to establish and 

spread.” 

APHIS recognizes that disturbed areas where tree removal activities may have taken place can be 

pathways for invasive weeds that already occur in the program area. The ALB Eradication 

Program, through compliance or cooperative agreements, stabilizes tree removal sites by planting 

grass to prevent erosion which also serves to minimize the introduction of invasive weeds. In wood 

lots, the program may opt not to plant grass and allow the natural seed back to sprout. 

Comment 21 

Several commenters express concerns over the use of imidicloprid in the Program, particularly in 

relation to the impacts to pollinators. One commenter wrote “I recommend waiting for the 

Presidential task force to finish its investigation and publish their recommendations as well as more 
G–8 Appendix G. Response to Comments 

research is published before the program uses neonicotinoids to treat potential ALB host trees, 

especially since the trees are in no way guaranteed to be safe from removal.” 

In the draft and final EIS, APHIS specifically addresses the risks to pollinators from ALB 

Eradication Program activities and conducted a risk assessment for imidicloprid to evaluate the 

risks to human health and ecological resources (including pollinators). The Program will continue 

to evaluate its use of chemicals to minimize the impact to nontarget species. 

Comment 22 

The DEIS states that in the 20 states of the USFS Northern Region, forests cover 42% of the land 

base and that vulnerable forest types (maple-beech-birch and elm-ash-cottonwood forests) 

collectively occupy about 45% of the forest area. We suggest noting that vulnerable forest types 

thus occupy almost 20% of the total land area. Severe damage to these forests as a result of ALB 

spread – especially in combination with the impact of other pests (see below) – will significantly 

damage all values provided by forests. Climate change is anticipated to bring about a shift away 

from ALB hosts such as maples to non-hosts such as oaks and pines. However, this shift does not 

promise a more healthy forest since these other tree types are vulnerable to other pests, e.g., 

European gypsy moth and winter moth. 

Comment noted. 

Comment 23 

In the beginning of the document when describing ALB damage, the authors describe ALB frass 

as saw dust like. Having seen ALB frass, I would suggest that it be described as long thin 

shavings. Sawdust frass more accurately describes frass from ants, leopard moth, and other 

insects. 

APHIS revised the description of frass to include thin shavings. 

Comment 24 

Several commenters pointed out that a non-host to ALB (yellow poplar, Lirodendron tulipifera) is 

included in the economic impact data. 

A correction in the final EIS clarifies that yellow poplar is not a host to ALB. 

Comment 25 

“Remove the infested part of the tree (pruning out damage) and leave the rest.” 

Usually when a tree is found infested with ALB, the tree already has multiple entry/exit wounds 

and it is unlikely all infested parts would be found for removal. 
Appendix G. Response to Comments G–9 

Comment 26 

“Cut an ALB “fire line” around regulated areas. This would involve removing all host material to 

create a 200m “ring” around the regulated area (the average distance ALB is thought to travel), 

through which the beetle would have to fly to reach more host material. The idea here being that 

most ALB will not reach the other side of the “fire line”.” 

In the ALB eradication program, the establishment of quarantine areas and the removal and/or 

treatment of trees within a certain distance of known infested trees is equivalent to the concept of a 

“fire line”. The quarantine boundaries and program activities are based on the biology, which 

includes the movement potential of the insect. 

Comment 27 

“If faced with funding restrictions, some targeted treatment programs may not actually complete a 

third consecutive year of treatment, as recommended. This was the case for the Worcester County, 

MA treatment program which treated high-risk host trees in Spring 2010 and Spring 2011 only. 

The third consecutive treatment year which should have occurred in Spring 2012 did not occur due 

to funding restrictions. This lack of treatment for a third consecutive year in Worcester County, 

MA limited the treatment program's maximum efficacy to prevent the spread of ALB, ultimately 

devaluing the first two treatment years in 2010 and 2011 (including all the wasted money, time, 

man hours, etc). For future ALB treatment endeavors to be successful, funding must be committed 

for the full consecutive three-year period. If not, the treatment program should be reconsidered or 

reconfigured.” 

Comment noted. 

Comment 28 

“Regarding imidacloprid and its potential effects on air quality, page 76 mentions the following: 

"The trunk injection method involves injecting imidacloprid directly into the tree trunk; this is 

unlikely to affect air quality due to minimal exposure of the insecticide to the ambient air." This 

statement may not be accurate when there are inclement weather conditions on treatment days or 

days prior to treatment. Trees that are already "saturated" with rain will not readily absorb injected 

imidacloprid like they will when not saturated. In this situation, some injected imidacloprid can 

and will ooze out of the drilled holes of the tree and seep onto the tree's root collar and/or perhaps 

the surrounding soil where it would remain until eventually being absorbed. Related to this, on 

page 81, the EIS specifically mentions that "No material may puddle or run offsite. " This also may 

not always be the case during inclement weather conditions. Due to the need to meet scientific and 

programmatic deadlines, treatment of high-risk host trees may not have the flexibility to always 

use ideal weather days. (In this context, a scientific deadline is referring to the life cycle of the 

ALB and the need to treat all necessary high-risk host trees prior to the first emergence of ALBs in 

the spring of a particular year). Inclement weather days may need to be part of the program's 

treatment schedule in 2010 and 2011 in Worcester County, MA; if these days had not been 

included in the schedule, there would not have been enough time to treat all of the high-risk host 

trees each spring before the beetles start emerging for the season.” 
G–10 Appendix G. Response to Comments 

APHIS recognizes that some material during tree injection may not move into the tree however the 

amount of material that would be likely to exit the injection holes would not be sufficient to impact 

air quality since imidacloprid is not considered an insecticide that is likely to volatilize into the 

atmosphere based on its chemical properties that are discussed in Appendix F. Imidacloprid 

applications are planned to occur during times that will maximize tree uptake which in turn 

increases efficacy. 

Comment 29 

One commenter points out other trees besides sugar maple are tapped for syrup. “Without taking 

steps to ensure that all potential syrup-producing tree species are tagged, the Program would 

inadvertently create more potential pathways and exposure risk than the draft EIS describes. With 

the potential for the demand for syrup products to grow (as noted on the bottom of page 50), more 

producers may enter the market or begin tapping their own trees recreationally, which may 

complicate outreach, education and notification efforts around the use of imidacloprid in treated 

trees. The draft EIS states that cumulative impacts to human health are negligible due to minimal 

risk of exposure, low frequency and duration of exposure in most situations. However, if a 

significant gap exists between the number of trees that are being treated and the tagging efforts, 

more cumulative exposure than anticipated may occur. It is essential that protocols are developed 

and used consistently in all regional treatment programs to ensure that every species of ALB host 

tree which has the potential to produce sap/syrup is included in the treatment efforts.” 

Comment noted. 

Comment 30 

“The draft EIS states that removal of infested trees will be done by the Program. However based on 

past experience, in some situations infested trees may be removed by landowners (e.g. for 

firewood) who may or may not follow proper techniques and procedures to minimize the spread of 

ALB. Education and outreach around the ALB should include strong discouragement in order to 

minimize the chance of individuals taking such actions. Precautions should include strict 

monitoring of infested trees after their identification, and appropriate actions taken to ensure only 

the Program handle removals.” 

APHIS cannot restrict homeowners from cutting down trees from their own properties but 

regulations are in place to restrict the movement of infested material. Both State and Federal law 

establish regulated areas around ALB infestations. The Federal regulation can be found in the 

Code of Federal Regulations (CFR), specifically in 7 CFR 301.51. Requirements for movement of 

regulated articles are outlined in those regulations. When an inspector has probable cause to 

believe a person is moving a regulated article interstate, the inspector is authorized to stop the 

person to determine whether a regulated article is present and to inspect it. The inspector may 

seize, quarantine, treat, destroy, dispose, or subject to other remedial measures infested articles. In 

addition, the State may impose restrictions on intrastate movement of regulated articles that are 

equivalent to those imposed on the interstate movement of regulated articles. 
Appendix G. Response to Comments G–11 

Comment 31 

“A few locations in the document explains that host material may leave the quarantine after 

being inspected, deemed uninfested, and with a permit. Later in the document on page 21 the 

author states “no method of visual survey...is completely effective...” If survey is not completely 

effective, how can APHIS take the chance of allowing regulated material to leave? What type of 

material is being allowed to leave? My suggestions to this would be to set guidelines on what 

may be considered allowed to leave the quarantine (ie. small nursery stock that has been grown 

from seed or plugs inside a greenhouse, small saplings less than 1-2 inches in diameter and 

shorter than 4-5 feet that can be easily and completely inspected for ALB, or must be a certain 

distance from known historical infested trees.)” 

The Code of Federal Regulations Title 7 §301.51- Subpart Asian Longhorned Beetle describes the 

conditions under which regulated articles may be permitted to leave a quarantine area. 

Comment 32 

“A question I had while reading the sections on erosion and water is, does APHIS consult with 

NRCS when doing removals in large areas where soil and water may be disturbed? It is a related 

USDA agency with the expertise in this subject. By working together and providing 

documentation to property owners and municipalities, it shows that APHIS is taking all steps in 

protecting the natural environment.” 

APHIS cooperates with other federal and state agencies, where appropriate, in a variety of its 

programs. This is the case with the ALB program as well and has included working with NRCS 

regarding site specific conditions. 

Comment 33 

“...why are the wood chips from large removal operations hauled away? Why not leave them 

scattered on site. A layer of wood chips would act as mulch and provide for erosion control, return 

nutrients to the soil, act as a barrier to invasive species, and support growth of the remaining 

understory and saplings. Another benefit of the woodchips is to offset soil compaction from the use 

of the heavy machinery during removals. If the chips are spread ahead of the machines they will 

have a layer of protection to work on. A study and article titled “Mechanical mastication thins 

Lake Tahoe forest with few adverse impacts” showed that leaving the chips resulted in minimal 

soil compaction and protected the soil from erosion and runoff. With today’s machinery and 

equipment I don’t see why this couldn’t physically be done, and believe that current practice of 

removing them is only beneficial to the removal company that can then sell the chips for even 

more profit.” 

APHIS will consider this practice in the ALB eradication program, adapting it to site specific 

needs. 
G–12 Appendix G. Response to Comments 

(This page is intentionally left blank.) 
Appendix H. Acronyms and Glossary H–1 

Appendix H. Acronyms and Glossary

A

Having a sudden onset, sharp rise, and short course 

Pleasing in appearance, relating to or dealing with the beautiful. 

Asian longhorned beetle

Rivers designated by the U.S. Environmental Protection Agency to 

receive special attention to further three objectives: natural resource 

and environmental protection, economic revitalization, and historic 

and cultural preservation. 

Animal and Plant Health Inspection Service; an agency within the U.S. 

Department of Agriculture. 

 The amount of pesticide product applied per unit area. 

Air Quality Index 

Of, in, or relating to water. 

There are 12 genera of host trees that APHIS regulates for ALB and are 

considered high-risk hosts: Acer (maple and box elder), Aesculus (horse 

chestnut and buckeye), Albizia (mimosa), Betula (birch), Cercidiphyllum 

(katsura tree), Fraxinus (ash), Koelreuteria (golden raintree), Platanus 

(sycamore and London planetree), Populus (poplar), Sorbus (mountain 

ash), Salix (willow), and Ulmus (elm). These trees are hosts because ALB 

can derive its food supply from them and complete its life cycle. 

B 

The area of a given section of land that is occupied by the cross-section of 

tree trunks and stems at their base. 

Black carbon; a particulate matter (PM) air pollutant formed by the 

incomplete combustion of fossil fuels, biofuels, and biomass.

Bird Conservation Region 

Butoxyethyl ester 

Acute 

Aesthetic 

ALB 

American 

Heritage 

River 

APHIS 

Application 

Rate 

AQI 

Aquatic 

Asian 

Longhorned 

Beetle 

High- Risk 

Host Tree 

Basal Area 

BC 

BCR 

BEE 
H–2 Appendix H. Acronyms and Glossary

A set of preventative measures used to protect soil and water quality from 

human disturbance. 

The number and variety of different organisms in the ecological complexes 

of which those organisms occur naturally; the relative abundance and 

frequency of biological organisms within ecosystems. 

The reduction of pest populations by means of living organisms introduced 

or supplemented by humans; utilizes competitors, parasites, predators, or 

sterile insects to reduce pest populations (also called biocontrol). 

Document stating the opinions of the U.S. Fish and Wildlife Service and the 

National Marine Fisheries Service as to whether a Federal action is likely to 

jeopardize the continued existence of a threatened or endangered species or 

result in the destruction or adverse modification of critical habitat. 

 C 

Centigrade or Celsius 

Clean Air Act 

A cylindrical layer of tissue in the stems and roots of many seed-bearing 

plants, consisting of cells that divide rapidly to form new layers of tissue. 

The uppermost layer in a forest, formed by the crowns of the trees. 

A substance capable of causing cancer in living tissue. 

Colony Collapse Disorder 

Council on Environmental Quality 

Code of Federal Regulations (U.S.) 

Methane

An adverse reaction(s) of a person or organism to ambient levels of toxic 

marked by long duration, by frequent recurrence over a long time, and often 

by slowly progressing seriousness. 

Marked by long duration, by frequent recurrence over a long time, and often 

by slowly progressing seriousness. 

Biodiversity

Biological

Control 

Best 

Management 

Practices 

Biological

Opinion 

oC

CAA

Cambium

Canopy 

Carcinogen/ 

Carcinogenic

CFR

CCD

CEQ 

CH4

Chemical 

Sensitivity 

Chronic
Appendix H. Acronyms and Glossary H–3 

Carbon Monoxide 

Carbon Dioxide 

The process of burning something. 

An assemblage of populations of plants, animals, bacteria, and fungi that live 

in an environment and interact with one another, forming a distinctive living 

system with its own composition, structure, environmental relations, 

development, and function; an association of interacting populations, usually 

defined by the nature of their interaction or the place in which they live. 

The ratio of the mass or volume of a solute to the mass or volume of the 

solution or solvent; the amount of active ingredient or herbicide equivalent 

in a quantity of diluent (e.g., expressed as lb/gal, ml/liter, etc.), or an amount 

of a substance in a specified amount of medium (e.g., air and water). 

A tree that bears cones and evergreen needlelike or scalelike leaves. 

State agriculture or forestry officials, contractors, and other entities 

collaborating with APHIS. 

Clean Water Act 

Coastal Zone Management Act

D 

Something that sheds after a period of time, such as a tree or a shrub that 

sheds its leaves annually. 

A route by which substances can enter the body through the skin. 

E 

Environmental Assessment; a concise public document which provides 

sufficient evidence and analysis for determining whether to prepare an EIS 

or Finding of No Significant Impact. It aids in compliance with the National 

Environmental Policy Act (NEPA) when no EIS is needed. 

A biological community of interacting organisms and their physical 

environment. 

The potential impact on an ecosystem from the release of toxic substances. 

 Community

CO2

Combustion

CO

Concentration

Cooperator

CWA

Conifer

CZMA

Dermal 

Absorption

Deciduous

EA

Ecosystem

Ecotoxicity
H–4 Appendix H. Acronyms and Glossary

Environmental Impact Statement; a document prepared by a Federal agency 

in which anticipated environmental effects of alternative planned courses of 

action are evaluated; a detailed written statement as required by section 

102(2)(C) of the National Environmental Policy Act (NEPA). 

The complete destruction or elimination of something. 

Executive Order 

U.S. Environmental Protection Agency 

Endangered Species Act 

 

F 

Fahrenheit 

An area of low-lying ground adjacent to a river, formed mainly of river 

sediments and subject to flooding. 

Plant leaves, collectively. 

Finding of no significant impact 

Federal Register 

Debris or excrement produced by insects. 

Forest Service; an agency of the U.S. Department of Agriculture. 

U.S. Federal Highway Administration 

Fish and Wildlife Service; an agency of the U.S. Department of Interior.

 

G 

Gram 

Greenhouse gas 

A gas that contributes to the greenhouse effect by absorbing infrared 

radiation. 

Gigaton

EIS

EO 

EPA

Eradication

ESA

oF 

FR

FS

FWHA

FWS 

Frass

Floodplain

Foliage

FONSI

g

GHG 

Gt

Greenhouse 

Gas (GHG) 
Appendix H. Acronyms and Glossary H–5 

H 

 

An angiosperm tree, especially as distinguished from a coniferous, or 

softwood, tree. It produces hard, compact wood or timber. Examples 

include oak, cherry, maple, or mahogany. 

The older, harder, nonliving, central wood of trees that is usually darker, 

denser, less permeable, and more durable than the surrounding sapwood. 

 

A substance used to kill plants. 

Human Health and Ecological Risk Assessment 

The science that studies the occurrence, circulation, distribution, and 

properties of the waters of the earth and its atmosphere. 

A state of altered reactivity in which the body reacts with an exaggerated 

immune response to what is perceived as a foreign substance. 

 

I 

To burn or reduce to ashes. 

The state of being invaded or overrun by pests or parasites. It can also refer 

to the actual organisms living on or within a host. 

The act of inhaling or breathing in. 

A substance used to kill insects. 

An animal lacking a backbone, such as an arthropod, mollusk, annelid, etc. 

International Plant Protection Convention 

International standards for phytosanitary measures 

 

K 

A species whose presence and role within an ecosystem has a 

disproportionate effect on other organisms within the ecosystem. 

Kilogram 

 

HHERA

Herbicide

Hardwood Tree

Heartwood 

Hydrology

Hypersensitivity

IPPC

ISPM

Insecticide

Infestation

Inhalation

Invertebrate

Incineration 

Kg

Keystone 

Species 
H–6 Appendix H. Acronyms and Glossary

L 

The active immature form of an insect, especially one that differs greatly 

from the adult and forms the stage between egg and pupa. 

Liter 

M 

Breakdown products of a chemical. 

Meter 

Microgram 

 

Milligram 

Memorandum of Understanding 

An agent, such as radiation or a chemical substance, that causes genetic 

mutation. 

N 

 

 Nitrous oxide 

National Ambient Air Quality Standards 

 A class of neuro-active insecticides chemically similar to nicotine. The 

insecticide imidacloprid is a neonicotinoid.

 A bird that breeds in Canada and the United States in the summer, and 

spends winter in Central America, South America, or the Caribbean 

Islands. 

The National Environmental Policy Act of 1969 and subsequent amendments. 

The ability of a drug or other agent to destroy or damage nervous tissue. 

Nanogram 

National Marine Fisheries Service 

Nitrate 

μ 

L 

mg

MOU

m

Metabolite

Larva

Mutagen/ 

Mutagenic

Neonicotinoid

Neotropica 

Bird 

NEPA

NAAQS

N2O

ng

NMFS

Neurotoxicity

NO3
Appendix H. Acronyms and Glossary H–7 

National Oceanic and Atmospheric Administration 

Nitrogen oxide 

No observable effect concentration 

No observable effect level; the highest dose levels at which there are no 

observable differences between the test and control populations. 

Notice of Intent 

O 

Nitrogen dioxide 

Ozone 

Office of Ocean and Coastal Resource Management 

A sudden increase in numbers of a harmful organism. 

To lay eggs. 

P 

A substance used for destroying insects or other organisms, including 

insecticides and herbicides. 

The ability of a substance to allow another substance to pass through it, 

especially the ability of a porous rock, sediment, or soil to transmit fluid 

through pores and cracks. 

Particulate matter 

Particulate matter that is less than 2.5 micrometers 

Particulate matter that is between 2.5 and 10 micrometers 

An insect or other organism that carries pollen from one flower to another.

Personal protective equipment 

parts per million 

NOAA

NOx

NOI

NOEL

NOEC

O2

O3

OCRM

Oviposit/ 

Oviposition

Outbreak

PM2.5

PM10

PPE

Pesticide

PM 

Permeability

Pollinator 

ppm
H–8 Appendix H. Acronyms and Glossary

Plant Protection and Quarantine 

A preventative measure. 

An insect in its inactive immature form between larva and adult. 

A larva becoming a pupa. 

 

Q 

A geographic area from which movement of regulated articles is restricted in 

order to prevent or reduce the human-assisted spread of a pest. APHIS 

establishes regulations that define the boundaries of quarantined areas. 

(Regulations for the ALB quarantined areas are located at 7 CFR § 301.51-

3.) 

R 

 

Articles that can harbor life stages of ALB. The regulated articles listed 

under the ALB quarantine (7 CFR § 301.51) include the beetle and all its life 

stages; firewood (all hardwood species, not restricted to ALB-host trees); 

green lumber, and other living, dead, cut, or fallen material, including 

nursery stock, logs, stumps, roots, branches, and debris from ALB host trees 

of 1⁄2 inch or more in diameter. 

Chronic reference dose; the term preferred by EPA to express acceptable 

daily intake for humans; an estimate (with uncertainty spanning perhaps an 

order of magnitude) of a daily exposure to the human population, 

including sensitive subgroups, that is likely to be without an appreciable risk 

of deleterious effects during a lifetime. 

Of or relating to wetlands adjacent to rivers and streams.

S 

The saltiness or dissolved salt content. 

An early and open process for determining the scope of issues to be 

addressed, and for identifying the significant issues related to a proposed 

action. 

Matter that settles to the bottom of a liquid. 

State Historic Preservation Office 

PPQ

Prophylactic

Pupa

Pupate

Quarantined 

Area/ 

Quarantine 

RfD

SHPO

Scoping

Regulated

Articles 

Riparian

Salinity 

Sediment
Appendix H. Acronyms and Glossary H–9 

The growing and cultivation of trees. 

State implementation plan 

Sterile insect technique 

Stream management zone 

Sulfur dioxide 

Sulfate 

Sulfur oxides 

All the organisms that spend a significant portion of their life cycle within a 

soil profile, or at the soil-litter interface. 

The process in which a stress applied to a soil causes densification as air is 

displaced from the pores between the soil grains. 

Removal of topsoil faster than the soil forming processes can replace it, due 

to natural, animal, and human activity. 

Not sufficient to cause death. 

An outlying part of a city or town; a smaller community adjacent to or 

within commuting distance of a city. 

An insecticide that is absorbed by the plant and moved throughout its tissues. 

T 

Threatened and endangered species 

Triethylamine salt 

An agent or factor that causes malformation of an embryo. 

Of, on, or relating to the earth. 

U 

A layer of vegetation beneath the main canopy of a forest. 

SIP

Systemic 

Insecticide

Silviculture

SIT

SO2

SO4

SOX

Soil Biota 

Soil Erosion 

Sublethal

Soil 

Compaction

Suburb/ 

Suburban

SMZ

T&E Species 

TEA

Teratogen/ 

Teratogenic 

Terrestrial

Understory
H–10 Appendix H. Acronyms and Glossary

Of, pertaining to, or designating a city or town; or living in a city. 

United States Code 

United State Department of Agriculture 

Ultraviolet 

V 

An animal with a backbone or spinal column, including mammals, birds, 

reptiles, amphibians, and fishes. 

Volatile organic compound 

Cause a substance to evaporate or disperse in vapor. 

W 

Wood packaging material 

U.S.C.

USDA

VOC 

WPM

Volatilize

UV

Urban

Vertebrate
Appendix I. References I–1 

I. References 

 Aizen, M. A. and P. Feinsinger. 1994. Forest fragmentation, pollination, and plant reproduction 

in a Chaco dry forest, Argentina. Ecology 75:330-351. 

Akbari, H., D. M. Kurn, S. E. Bretz, and J. W. Hanford. 1997. Peak power and cooling energy 

savings of shade trees. Energy and Buildings 25. 

Alaux, C., J.-L. Brunet, C. Dussaubat, F. Mondet, S. Tchamitchan, M. Cousin, J. Brillard, A. 

Baldy, L. P. Belzunces, and Y. Le Conte. 2010. Interactions between Nosema 

microspores and a neonicotinoid weaken honeybees (Apis mellifera). Environmental 

Microbiology 12:774-782. 

Allen-Wardell, G., P. Bernhardt, R. Bitner, A. Burquez, S. Buchmann, J. Cane, P. A. Cox, V. 

Dalton, P. Feinsinger, M. Ingram, D. Inouye, C. E. Jones, K. Kennedy, P. Kevan, H. 

Koopowitz, R. Medellin, S. Medellin-Morales, P. Nabhan, B. Pavlik, V. Tepedino, P. 

Torchio, and S. Walker. 1998. The potential consequences of pollinator declines on the 

conservation of biodiversity and stability of food crop yields. Conservation Biology 12:8- 

17. 

Anatra-Cordone, M. and P. Durkin. 2005. Imidacloprid--Human Health and Ecological Risk 

Assessment--Final Report. United States Department of Agriculture, Forest Service. 

Anderson, L. M. and H. K. Cordell. 1988. Influence of trees on residential property values in 

Athens, Georgia (U. S. A.): A survey based on actual sales prices. Landscape and Urban 

Planning 15:153-164. 

APHIS. 2003. Importation of Solid Wood Packing Material, Final Environmental Impact 

Statement, August 2003. United States Department of Agriculture, Animal and Plant 

Health Inspection Service. 

APHIS. 2007a. Asian Longhorned Beetle: Questions and Answers (available at: 

http://www.aphis.usda.gov/publications/plant_health/content/printable_version/faq_alb_0 

7.pdf). United States Department of Agriculture, Animal and Plant Health Inspection 

Service, Plant Protection and Quarantine. 

APHIS. 2007b. Importation of Solid Wood Packing Material, Supplemental to the Final 

Environmental Impact Statement, October 2007. United States Department of 

Agriculture, Animal and Plant Health Inspection Service. 

APHIS. 2008. Asian longhorned beetle: Annotated host list. Page 2. United States Department of 

Agriculture, Animal and Plant Health Inspection Service, Plant Protection and 

Quarantine, Otis Pest Survey Detection and Exclusion Laboratory. 

APHIS. 2009. USDA–APHIS Asian Longhorned Beetle Eradication Program— Decision 

Support for the APHIS Management Team. Page 97 pp. United States Department of 

Agriculture, Animal and Plant Health Inspection Service. 

APHIS. 2013a. 2010-2012 Environmental Monitoring Report, Asian Longhorned Beetle 

Cooperative Eradication Program: For the Eradication Effort in Worcester and Boston, 

Massachusetts. United States Department of Agriculture, Animal and Plant Health 

Inspection Service, Plant Protection and Quarantine. 

APHIS. 2013b. Asian Longhorned Beetle Cooperative Eradication Program in Clermont County, 

Ohio. May 2013 United States Department of Agriculture, Animal and Plant Health 

Inspection Service, Riverdale, MD. 
I–2 Appendix I. References

Aust, W. M. and C. R. Blinn. 2004. Forestry best management practices for timber harvesting 

and site preparation in the eastern United States: An overview of water quality and 

productivity research during the past 20 years (1982-2002). Water, Air, and Soil 

Pollution: Focus 4:5-35. 

Ballard, T. M. 2000. Impacts of forest management on northern forest soils. Forest Ecology and 

Management 133:37-42. 

Bancroft, J. and M. T. Smith. 2005. Dispersal and influences on movement for Anoplophora 

glabripennis calculated from individual mark-recapture. Entomologia Experimentalis et 

Applicata 116:83-92. 

Barrios, E. 2007. Soil biota, ecosystem services and land productivity. Ecological Economics 

64:269-285. 

Batra, S. W. T. 1985. Red Maple (Acer rubrum L.), an Important Early Spring Food Resource 

for Honey Bees and Other Insects. Journal of the Kansas Entomological Society 58:169- 

172. 

Beckett, K. P., P. H. Freer-Smith, and G. Taylor. 1998. Urban woodlands: their role in reducing 

the effects of particulate pollution. Environmental Pollution 99:347-360. 

Bell, J. F., J. S. Wilson, and G. C. Liu. 2008. Neighborhood greenness and 2-year changes in 

body mass index of children and youth. Am J Prev Med 35:547-553. 

Bell, J. L. and R. C. Whitmore. 2000. Bird nesting ecology in a forest defoliated by gypsy moths. 

Wilson Bulletin 112:524-531. 

Bell, J. N. B. and M. Treshow. 2002. Air pollution and plant life. 2nd edition. John Wiley and 

Sons, New York. 

Bréda, N., R. Huc, A. Granier, and E. Dreyer. 2006. Temperate forest trees and stands under 

severe drought: a review of ecophysiological responses, adaptation processes and longterm consequences. Ann. For. Sci. 63:625-644. 

Bulman, L. S., M. O. Kimberley, and P. D. Gadgil. 1999. Estimation of the efficiency of pest 

detection surveys. New Zealand Journal of Forestry Science 29:102-115. 

Bush, E. B., D. G. Neary, C. K. McMahon, and J. W. Taylor Jr. 1987a. Suitability of hardwoods 

treated with phenoxy and pyridine herbicides for use as firewood. Archives of 

Environmental Contamination and Toxicology 16:333-341. 

Bush, P. B., J. W. Taylor, C. K. McMahon, and D. G. Neary. 1987b. Residues of lindane and 

chlorpyrifos in firewood and woodsmoke. J. Entomol. Sci. 22:131-139. 

Bush, R. J., J. J. Bejune, B. G. Hansen, and P. A. Araman. 2002. Trends in the use of materials 

for pallets and other factors affecting the demand for hardwood products.in 30th 

Hardwood Symposium. 

Cao, S., T. Tian, L. Chen, X. Dong, X. Yu, and G. Wang. 2010. Damage Caused to the 

Environment by Reforestation Policies in Arid and Semi-Arid Areas of China. AMBIO: 

A Journal of the Human Environment 39:279-283. 

Cavey, J. F. 1998. Solid wood packing material from China. Initial pest risk assessment on 

certain wood boring beetles known to be associated with cargo shipments: Asian 

longhorned beetle (Anoplophora glabripennis), Ceresium, Monochamus and 

Hesperophanes. National Identification Services - SS - PPQ - APHIS - USDA, Riverdale, 

MD. 

Chalfoun, A. D., F. R. I. Thompson, and M. J. Ratnaswamy. 2002. Nest predators and 

fragmentation: A review and meta-analysis. Conservation Biology 16:306-318. 
Appendix I. References I–3 

Charlier, R., P. Hornych, M. Sršen, A. Hermansson, G. Bjarnason, S. Erlingsson, and P. Pavšič. 

2009. Water influence on bearing capacity and pavement performance: Field 

observations. Pages 175-192 in Geotechnical, Geological and Earthquake Engineering, 

Volume 5. 

Chen, X. D., E. Culbert, V. Hebert, and J. D. Stark. 2010. Mixture effects of the nonylphenyl 

polyethoxylate, R-11 and the insecticide, imidacloprid on population growth rate and 

other parameters of the crustacean, Ceriodaphnia dubia. Ecotoxicology and 

Environmental Safety 73:132-137. 

Cook, M. 1990. Soil and water quality. AG-North Carolina Agricultural Extension Service, 

North Carolina State University. 

Coulson, R. N., M. A. Pinto, M. D. Tchakerian, K. A. Baum, W. L. Rubink, and J. S. Johnston. 

2005. Feral honey bees in pine forest landscapes of east Texas. Forest Ecology and 

Management 215:91-102. 

Coulson, R. N. and F. M. Stephen. 2006. Impacts of insects in forest landscapes: Implications for 

forest health management. Pages 101-125 Invasive Forest Insects, Introduced Forest 

Trees, and Altered Ecosystems. Springer. 

Cox, G. W. 1999. Eastern forests: The dark side of forest biodiversity. Pages 96-109 Alien 

species in North America and Hawaii: Impacts on natural ecosystems. Island Press. 

Cunningham, W. P., B. W. Saigo, and M. A. Cunningham. 2001. Environmental science: A 

global concern. McGraw-Hill Boston, MA. 

Dayton, P. K. 1972. Toward an understanding of community resilience and the potential effects 

of enrichments to the benthos at McMurdo Sound, Antarctica. Pages 81-96 in 

Proceedings of the colloquium on conservation problems in Antarctica. Allen Press 

Lawrence, Kansas, USA. 

Desneux, N., A. Decourtye, and J.-M. Delpuech. 2007. The sublethal effects of pesticides on 

beneficial arthropods. Annual Review of Entomology 52:81-106. 

Dodds, K. J., H. M. Hull-Sanders, N. W. Siegert, and M. J. Bohne. 2014. Colonization of three 

maple species by Asian longhorned beetle, Anoplophora glabripennis, in two mixedhardwood forest stands. Insects 5:105-119. 

Dombrow, J. M. R. and C. F. Sirmans. 2000. The market value of mature trees in single-family 

housing markets. The Appraisal Journal 68:39-43. 

Donovan, G. H. and D. T. Butry. 2010. Trees in the city: Valuing street trees in Portland, 

Oregon. Landscape and Urban Planning 94:77-83. 

Donovan, G. H., D. T. Butry, Y. L. Michael, J. P. Prestemon, A. M. Liebhold, D. Gatziolis, and 

M. Y. Mao. 2013. The relationship between trees and human health. American Journal of 

Preventive Medicine 44:139-145. 

Donovan, G. H., Y. L. Michael, D. T. Butry, A. D. Sullivan, and J. M. Chase. 2011. Urban trees 

and the risk of poor birth outcomes. Health & Place 17:390-393. 

Donovan, T. M., F. R. I. Thompson, J. Faaborg, and J. R. Probst. 1995. Reproductive success of 

migratory birds in habitat sources and sinks. Conservation Biology 9:1380-1395. 

Doran, J. W. and T. B. Parkin. 1994. Defining and assessing soil quality. Special Publication 

35.in J. W. Doran, D. C. Coleman, D. F. Bezdicek, and B. A. Stewart, editors. Defining 

soil quality for a sustainable environment. Soil Science Society of America, Madison, 

WI:. 

Dosskey, M. G., P. Vidon, N. P. Gurwick, C. J. Allan, T. P. Duval, and R. Lowrance. 2010. The 

role of riparian vegetation in protection and improving chemical water quality in streams. 

Journal of the American Water Resources Association 46:261-277. 
I–4 Appendix I. References

Doupé, R. G., J. Mitchell, M. J. Knott, A. M. Davis, and A. J. Lymbery. 2010. Efficacy of 

exclusion fencing to protect ephemeral floodplain lagoon habitats from feral pigs (Sus 

scrofa). Wetlands Ecology and Management 18:69-78. 

DPA. 2013. Maine Office of Tourism, Visitor Tracker Research, 2012 Calendar Year Annual 

Report. DPA, Kennebunk, ME. 

Durkin, P. 2003. Triclopyr: Revised Human Health and Ecological Risk Assessments - Final 

Report. Page 264. United States Department of Agriculture, Forest Service. 

Durkin, P. and M. Follansbee. 2004. Imazapyr - Human Health and Ecological Risk Assessment 

- Final Report. United States Department of Agriculture, Forest Service. 

Dwyer, J., E. G. McPherson, H. W. Schroeder, and R. A. Rowntree. 1992. Assessing the benefits 

and costs of the urban forest. Journal of Arboriculture 18:227-234. 

Ellison, A. M., M. S. Bank, B. D. Clinton, E. A. Colburn, K. Elliot, C. R. Ford, D. R. Foster, B. 

D. Kloeppel, J. D. Knoepp, G. M. Lovett, J. Mohan, D. A. Orwig, N. L. Rodenhouse, W. 

V. Sobczak, K. A. Stinson, J. K. Stone, C. M. Swan, J. Thompson, B. Von Holle, and J. 

R. Webster. 2005. Loss of foundation species: Consequences for the structure and 

dynamics of forested ecosystems. Frontiers in Ecology and the Environment 3:479-486. 

Englin, J. E., J. M. McDonald, and K. Moeltner. 2006. Valuing ancient forest ecosystems: An 

analysis of backcountry hiking in Jasper National Park. Ecological Economics 57:665- 

678. 

EPA. 1998. Reregistration eligibility decision (RED): Triclopyr (EPA 738-R-98-011). Page 285. 

United States Environmental Protection Agency. 

EPA. 2008. Reducing Urban Heat Islands: Compendium of Strategies. Available at: 

http://www.epa.gov/heatisland/resources/compendium.htm. United States Environmental 

Protection Agency (EPA). 

EPA. 2012a. National Ambient Air Quality Standards (NAAQS). 

EPA. 2012b. Our Nation’s Air: Status and Trends through 2010. EPA-454/R-12-001. Last 

accessed May 21, 2014 at: http://epa.gov/airtrends/2011/report/fullreport.pdf. U.S. 

Environmental Protection Agency, Office of Air Quality Planning and Standards. 

EPA. 2013a. The Green Book Nonattainment Areas for Criteria Pollutants. Last accessed April 

24, 2014 at: http://www.epa.gov/airquality/greenbook/index.html. 

EPA. 2013b. Water Quality Assessment and TMDL Information: National Summary of State 

Information. United States Environmental Protection Agency. 

EPA. 2013c. Water: Estuaries and Coastal Watersheds, National Estuary Program Overview. 

June 10, 2013. Available at: http://water.epa.gov/type/oceb/nep/, last accessed January 

10, 2014. United States Environmental Protection Agency (EPA). 

EPA. 2014a. Ecological Fate and Effects Division (EFED) Pesticide Ecotoxicity Database. 

United States Environmental Protection Agency, Office of Pesticide Programs. 

EPA. 2014b. Inventory of U.S. Greenhouse Gas Emissions and Sinks: 1990-2012 (EPA 430-R14-003). United States Environmental Protection Agency, Washington, D.C. 

EPA. 2014c. Watershed Assessment, Tracking and Environmental Results (Available at: 

http://iaspub.epa.gov/waters10/attains_nation_cy.control?p_report_type=T#imp_water_b 

y_state). United States Environmental Protection Agency. 

EPA. n.d. Volatile organic compounds emissions. United States Environmental Protection 

Agency. 

EPPO. 1999. EPPO data sheets on quarantine pests: Anoplophora glabripennis. European Plant 

Protection Organization. 

Everson-Rose, S. A. and T. T. Lewis. 2005. Psychosocial factors and cardiovascular diseases. 
Appendix I. References I–5 

Annual Review of Public Health 26:469-500. 

FAO. 2002 International Standards for Phytosanitary Measures (ISPM) No. 15. Guidelines for 

Regulating Wood Packaging Material in International Trade. 

Farrell, M. L. and B. F. Chabot. 2012. Assessing the growth potential and economic impact of 

the U.S. maple syrup industry. Journal of Agriculture, Food Systems, and Community 

Development 2:11-27. FAS. 2009. U. S. trade internet system. United States 

Department of Agriculture, Foreign Agriculture Service. 

FS. 2009a. Susceptibility potential for Anoplophora glabripennis (Asian longhorned beetle). 

United States Department of Agriculture, Forest Service. 

FS. 2009b. Timber Product Output (TPO) reports. United States Department of Agriculture, 

Forest Service. 

FS. 2014. Forest Inventory and Analysis Glossary. Last accessed July 1, 2014 at 

http://www.nrs.fs.fed.us/fia/data-tools/state-reports/glossary/default.asp. United States 

Department of Agriculture, Forest Service. 

FWS. 2006. A blueprint for the design and delivery of bird conservation in the Atlantic northern 

forest. Page 346. United States Department of the Interior, Fish and Wildlife Service. 

FWS. 2007. Indiana Bat (Myotis sodalis) Draft Recovery Plan: First Revision. United States 

Department of the Interior, U.S. Fish and Wildlife Service, Great Lakes-Big Rivers 

Region - Region 3, Fort Snelling, Minnesota. 

Garrison, G. A., A. J. Bjugstad, D. A. Duncan, M. E. Lewis, and D. R. Smith. 1977. Vegetation 

and environmental features of forest and range ecosystems. Page 475 in U.-F. A. 

Handbook, editor. U. S. Government Printing Office. 

Gill, R. J., O. Ramos-Rodriguez, and N. E. Raine. 2012. Combined pesticide exposure severely 

affects individual- and colony-level traits in bees. Nature 491:105-108. 

Gilliom, R. J., J. E. Barbash, C. G. Crawford, P. A. Hamilton, J. D. Martin, N. Nakagaki, L. H. 

Nowell, J. C. Scott, P. E. Stackelberg, G. P. Thelin, and D. M. Wolock. 2007. The 

Quality of our Nation's Waters: Pesticides in the Nation's Streams and Ground Water, 

1992-2001. Circular 1291. United States Department of the Interior, U.S. Geological 

Survey. 

Grahn, P. and U. A. Stigsdotter. 2003. Landscape planning and stress. Urban Forestry & Urban 

Greening 2:1-18. 

Grant, R. H., G. M. Heisler, and W. Gao. 2002. Estimation of pedestrian level UV exposure 

under trees. Photochemistry and Photobiology 75:369-376. 

GTIS. 2013. Global Trade Atlas. Global Trade Information Services (GTIS), Inc. 

Guite, H. F., C. Clark, and G. Ackrill. 2006. The impact of the physical and urban environment 

on mental wellbeing. Public Health 120:1117-1126. 

Haack, R. A., L. S. Bauer, R. Gao, J. J. McCarthy, D. L. Miller, T. R. Petrice, and T. M. Poland. 

2006. Anoplophora glabripennis within-tree distribution, seasonal development, and host 

suitability in China and Chicago. The Great Lakes Entomologist 39:169-183. 

Haack, R. A., F. Hérard, J. Sun, and J. J. Turgeon. 2010. Managing invasive populations of 

Asian longhorned beetle and citrus longhorned beetle: A worldwide perspective. Annual 

Review of Entomology 55:521-546. 

Haack, R. A., K. R. Law, V. C. Mastro, S. H. Ossenburgen, and B. J. Raimo. 1997. New York's 

battle with the Asian long-horned beetle. Journal of Forestry 95:11-15. 

Hajek, A. E. 2010. Asian Longhorned Beetle: Ecology and Control. Pages 21-24 in D. Pimental, 

editor. Encyclopedia of Pest Management. CRC Press. 
I–6 Appendix I. References

Hartig, T., G. W. Evans, L. D. Jamner, D. S. Davis, and T. Gärling. 2003. Tracking restoration in 

natural and urban field settings. Journal of Environmental Psychology 23:109-123. 

HCN and DAC. 2004. Nature and Health: The influence of nature on social, psychological and 

physical well-being. The Hague: Health Council of the Netherlands and RMNO; 

publication no. 2004/09E; RMNO publication nr A02ae. Health Council of the 

Netherlands (HCN) and Dutch Advisory Council (DAC) for Research on Spatial 

Planning Nature and the Environment. 

Heisler, G. and R. H. Grant. 2000. Ultraviolet Radiation, Human Health, and the Urban Forest. 

Heisler, G. M. 2010. UV Radiation in Global Climate Change: Measurements, Modeling and 

Effects on Ecosystems. Page 544 pp. in W. Gao, D. L. Schmoldt, and J. R. Slusser, 

editors. Urban Forest Influences on Exposure to UV Radiation and Potential 

Consequences for Human Health (Chapter 12). Tsinghua University Press, Beijing. 

Heisler, G. M., R. H. Grant, and W. Gao. 2003. Individual- and scattered-tree influences on 

ultraviolet irradiance. Agricultural and Forest Meteorology 120:113-126. 

Henley, W. F., M. A. Patterson, R. J. Neves, and A. D. Lemly. 2000. Effects of sedimentation 

and turbidity on lotic food webs: A concise review for natural resource managers. 

Reviews in Fisheries Science 8:125-139. 

Hladik, M. L., D. W. Kolpin, and K. M. Kuivila. 2014. Widespread occurrence of neonicotinoid 

insecticides in streams in a high corn and soybean producing region, USA. 

Environmental Pollution 193:189-196. 

Holmes, T. P., E. A. Murphy, and K. P. Bell. 2006. Exotic forest insects and residential property 

values. Agricultural and Resource Economics Review 35:155-166. 

Holmstrup, M., A.-M. Bindesbøl, G. J. Oostingh, A. Duschl, V. Scheil, H.-R. Köhler, S. 

Loureiro, A. M. V. M. Soares, A. L. G. Ferreira, C. Kienle, A. Gerhardt, R. Laskowski, 

P. E. Kramarz, M. Bayley, C. Svendsen, and D. J. Spurgeon. 2010. Interactions between 

effects of environmental chemicals and natural stressors: A review. Science of The Total 

Environment 408. 

Hornbeck, J. W., M. B. Adams, E. S. Corbett, E. S. Verry, and J. A. Lynch. 1993. Long-term 

impacts of forest treatments on water yield: a summary for northeastern USA. Journal of 

Hydrology 150:323-344. 

Horsley, S. B., R. P. Long, S. W. Bailey, R. A. Hallett, and P. M. Wargo. 2002. Health of eastern 

North American sugar maple forests and factors affecting decline. Northern Journal of 

Applied Forestry 19:34-44. 

Hu, J., S. Angeli, S. Schuetz, Y. Luo, and A. E. Hajek. 2009. Ecology and management of exotic 

and endemic Asian longhorned beetle Anoplophora glabripennis. Agricultural and Forest 

Entomology 11:359-375. 

Huang, Y. J., H. Akbari, H. Taha, and A. H. Rosenfeld. 1987. The potential of vegetation in 

reducing summer cooling loads in residential buildings. Journal of Climate and Applied 

Meteorology 26:1103–1116. 

Humpel, N., N. Owen, and E. Leslie. 2002. Environmental factors associated with adults’ 

participation in physical activity. Am. J. Prev. Med. 22:188- 199. 

Iverson, L. R. and A. M. Prasad. 1998. Predicting abundance of 80 tree species following climate 

change in the eastern United States. Ecological Monographs 68:465-485. 

Iverson, L. R., A. M. Prasad, S. N. Matthews, and M. Peters. 2008. Estimating potential habitat 

for 134 eastern US tree species under six climate scenarios. Forest Ecology and 

Management 254:390-406. 
Appendix I. References I–7 

Jacobson, M. G., R. C. Ready, and C. D. Canham. 2012. Potential ecological and economic 

impacts of a regional outbreak of the Asian longhorned beetle. Final report for USDAAPHIS Award 10-8100-1442-CA. The Pennsylvania State University, State College, PA. 

Johnson, J. 2012. The role of pesticides on honey bee health and hive maintenance with an 

emphasis on the neonicotinoid, imidacloprid. University of Maryland, Baltimore, 

Maryland. 

Jones, K. L., G. C. Poole, J. L. Meyer, W. Bumback, and E. A. Kramer. 2006. Quantifying 

expected ecological response to natural resource legislation: a case study of riparian 

buffers, aquatic habitat, and trout populations. [Online.] Available: 

http://www.ecologyandsociety.org/vol11/iss2/art15/. Ecology and Society 11. 

Kane, B. and J. Kirwan. 2009. Value, Benefits, and Costs of Urban Trees. Virginia Cooperative 

Extension Publication 420–181. Last accessed March 31, 2014: 

http://www.pubs.ext.vt.edu/420/420-181/420-181_pdf.pdf

Karlen, D., M. Mausbach, J. Doran, R. Cline, R. Harris, and G. Schuman. 1997. Soil quality: A 

concept, definition, and framework for evaluation (a guest editorial). Soil Science Society 

of America Journal 61:4-10. 

Kaushal, S. S., G. E. Likens, N. A. Jaworski, M. L. Pace, A. M. Sides, D. Seekell, K. T. Belt, D. 

H. Secor, and R. L. Wingate. 2010. Rising stream and river temperatures in the United 

States. Frontiers in Ecology and the Environment 8:461-466. 

Keena, M. 2002. Anoplophora glabripennis (Coleoptera: Cerambycidae) fecundity and longevity 

under laboratory conditions: comparison of populations from New York and Illinois on 

Acer saccharum. Environmental Entomology 3:490-498. 

Keena, M. 2006. Effects of temperature on Anoplophora glabripennis (Coleoptera: 

Cerambycidae) adult survival, reproduction, and egg hatch. Environmental Entomology 

35:912-921. 

Keena, M. 2008. Phenology of the Asian longhorned beetle under simulated annual 

environmental fluctuations. United States Department of Agriculture. 

Kim, H. and J. Bernstein. 2009. Air pollution and allergic disease. Current Allergy and Asthma 

Reports 9:128-133. 

Kizlinski, M. L., D. A. Orwig, R. C. Cobb, and D. R. Foster. 2002. Direct and indirect ecosystem 

consequences of an invasive pest on forest dominated by eastern hemlock Jounal of 

Biogeography 29:1489-1503. 

Klein, A. M., B. E. Vaissière, J. H. Cane, I. Ingolf Steffan-Dewenter, S. A. Cunningham, C. 

Kremen, and T. Tscharntke. 2007. Importance of pollinators in changing landscapes for 

world crops. Proceedings of the Royal Society B-Biological Sciences 274:303-313. 

Klotzback, J. and P. Durkin. 2004. Metsulfuron Methyl - Human Health and Ecological Risk 

Assessment - Final Report (SERA TR 04-43-17-01c). United States Department of 

Agriculture, Forest Service. 

Knoepp, J. D. and B. D. Clinton. 2009. Riparian zones in southern Appalachian headwater 

catchments: Carbon and nitrogen responses to forest cutting. Forest Ecology and 

Management 258:2282-2293. 

Koide, R. T. and T. Wu. 2003. Ectomycorrhizas and retarded decomposition in a Pinus resinosa

plantation. New Phytologist 158:401-407. 

Kraemer, M. E. and F. D. Favi. 2005. Flower phenology and pollen choice of Osmia lignaria 

(Hymenoptera: Megachilidae) in Central Virginia. Environmental Entomology 34:1593- 

1605. 
I–8 Appendix I. References

Kremen, C., N. M. Williams, M. A. Aizen, B. Gemmill-Herren, G. LeBuhn, R. Minckley, L. 

Packer, S. G. Potts, T. a. Roulston, I. Steffan-Dewenter, D. P. Vázquez, R. Winfree, L. 

Adams, E. E. Crone, S. S. Greenleaf, T. H. Keitt, A.-M. Klein, J. Regetz, and T. H. 

Ricketts. 2007. Pollination and other ecosystem services produced by mobile organisms: 

a conceptual framework for the effects of land-use change. Ecology Letters 10:299-314. 

Kreutzweiser, D., K. Gooda, D. Chartranda, T. Scarrb, and D. Thompson. 2007. Non-target 

effects on aquatic decomposer organisms of imidacloprid as a systemic insecticide to 

control emerald ash borer in riparian trees. Ecotoxicology and Environmental Safety 

68:315-325. 

Kreutzweiser, D. P., K. P. Good, D. T. Chartrand, T. A. Scarr, and D. G. Thompson. 2008a. Are 

leaves that fall from imidacloprid-treated maple trees to control Asian longhorned beetles 

toxic to non-target decomposer organisms? Journal of Environmental Quality 37:639- 

646. 

Kreutzweiser, D. P., K. P. Good, D. T. Chartrand, T. A. Scarr, and D. G. Thompson. 2008b. 

Toxicity of the systemic insecticide, Imidacloprid, to forest stream insects and microbial 

communities. Bull. Environ. Contam. Toxicol. 80:211-214. 

Kreutzweiser, D. P., D. G. Thompson, and T. A. Scarr. 2009. Imidacloprid in leaves from 

systemically treated trees may inhibit litter breakdown by non-target invertebrates. 

Ecotoxicology and Environmental Safety 72:1053-1057. 

Kuo, F. and W. Sullivan. 2001a. Aggression and violence in the inner city : Effects of 

environment via mental fatigue. Environment and Behavior 33:543-571. 

Kuo, F. E. and W. C. Sullivan. 2001b. Environment and crime in the inner city: does vegetation 

reduce crime? Environ Behav 33:343-367. 

Langmann, B., B. Duncan, C. Textor, J. Trentmann, and G. R. van der Werf. 2009. Vegetation 

fire emissions and their impact on air pollution and climate. Atmospheric Environment 

43:107-116. 

Laskowski, R., A. J. Bednarska, P. E. Kramarz, S. Loureiro, V. Scheil, J. Kudłek, and M. 

Holmstrup. 2010. Interactions between toxic chemicals and natural environmental factors 

— A meta-analysis and case studies. Science of The Total Environment 408:3763-3774. 

Lee, J., Park, B. J., Y. Tsunetsugu, T. Kagawa, and Y. Miyazaki. 2009. Restorative effects of 

viewing real forest landscapes, based on a comparison with urban landscapes. 

Scandinavian Journal of Forest Research 24:227-234. 

Lee, P., C. Smyth, and S. Boutin. 2004. Quantitative review of riparian buffer width guidelines 

from Canada and the United States. Journal of Environmental Management 70:165-180. 

Lingafelter, S. W. and E. R. Hoebeke. 2002. Revision of Anoplophora (Coleoptera: 

Cerambycidae). The Entomological Society of Washington, Washington, D.C. 

Logan, J. A., J. Regniere, and J. A. Powell. 2003. Assessing the impacts of global warming on 

forest pest dynamics. Frontiers in Ecology and the Environment 1:130-137. 

Loureiro, S., C. Svendsen, A. Ferreira, L. G., C. Pinheiro, F. Ribeiro, and A. M. V. M. Soares. 

2010. Toxicity of three binary mixtures to Daphnia manga: Comparing chemical modes 

of action and deviations from conceptual models. Environmental Toxicology and 

Chemistry 29:1716-1726. 

Lovasi, G. S., J. W. Quinn, K. M. Neckerman, M. S. Perzanowski, and A. Rundle. 2008. 

Children living in areas with more street trees have lower prevalence of asthma. J 

Epidemiol Community Health 62:647-649. 
Appendix I. References I–9 

Lovett, G. M., C. D. Canham, M. A. Arthur, K. C. Weathers, and R. D. Fitzhugh. 2006. Forest 

ecosystem responses to exotic pests and pathogens in eastern North America. Bioscience 

56:395-405. 

Lovett, G. M., L. M. Christenson, P. M. Groffman, C. G. Jones, J. E. Hart, and M. J. Mitchell. 

2002. Insect defoliation and nitrogen cycling in forests. Bioscience 52:335-341. 

Lovett, G. M. and M. J. Mitchell. 2004. Sugar maple and nitrogen cycling in the forests of 

eastern North America. Frontiers in Ecology and the Environment 2:81-88. 

Lovett, G. M. and H. Rueth. 1999. Soil nitrogen transformation in beech and maple stands along 

a nitrogen deposition gradient. Ecological Applications 9:1330-1344. 

Lu, C., K. M. Warchol, and R. A. Callahan. 2014. Sub-lethal exposure to neonicotinoids 

impaired honey bees winterization before proceeding to colony collapse disorder. 

Bulletin of Insectology 67:125-130. 

Lucas, S. R. and T. A. E. Platts-Mills. 2005. Physical activity and exercise in asthma: Relevance 

to etiology and treatment. J. Allergy Clin. Immunol. 115:928-934. 

McDaniel, P. 2006. The twelve soil orders: soil taxonomy (Last accessed June 30, 2014 at 

http://www.cals.uidaho.edu/soilorders/). University of Idaho, College of Agricultural and 

Life Sciences. 

McMahon, C. K., H. B. Clements, P. B. Bush, D. G. Neary, and J. W. Taylor. 1985. Pesticides 

released from burning treated wood. Pages 145-152 in Eighth Conference on Fire and 

Forest Meteorology, Detroit, Michigan. 

McPherson, E. G. 2005. Tress with benefits. American Nurseryman 201:34-40. 

McPherson, E. G. and J. R. Simpson. 1999. Carbon dioxide reduction through urban forestry: 

Guidelines for professional and volunteer tree planters. Gen. Tech. Rep. PSWGTR-171. 

Albany, CA: . United States Department of Agriculture, Forest Service, Pacific 

Southwest Research Station. 

McPherson, G. and J. R. Simpson. 1995. Shade trees as a demand-side resource. Last accessed 

July 2, 2014 at 

http://treebenefits.terrasummit.com/Documents/Air_Quality/Shade%20Trees%20as%20a

%20Demand-Side%20Resource.pdf. Home Energy Magazine Online. 

Mehrhoff, L. A. 1989. The dynamics of declining populations of an endangered orchid, Isoria 

medoloides. Ecology 70:783-786. 

Meiners, S. J., S. T. A. Pickett, and M. L. Cadenasso. 2002. Exotic plant invasions over 40 years 

of old field successions: Community patterns and associations. Ecography 25:215-223. 

Morales, D., B. N. Boyce, and R. J. Favretti. 1976. The contribution of trees to residential 

property value: Manchester, Connecticut. Valuation October/November:26-43. 

Mroz, G., M. Jurgensen, and D. Frederick. 1985. Soil nutrient changes following whole tree 

harvesting on three northern hardwood sites. Soil Science Society of America Journal 

49:1552-1557. 

Naiman, R. J. and H. Decamps. 1997. The ecology of interfaces: Riparian zones. Annual Review 

of Ecology and Systematics 28:621-658. 

NASS. 2014. Maple syrup production. United States Department of Agriculture, National 

Agricultural Statistics Service, Harrisburg, PA. 

NCFS. n.d. Riparian and Wetland Tree Planting Pocket Guide for North Carolina. Second 

Edition. Last accessed May 6, 2014: 

http://ncforestservice.gov/publications/WQ0206.pdf. North Carolina Forest Service. 
I–10 Appendix I. References

NHDES. 2008. A field guide to common aquatic and riparian plants of New Hampshire. Second 

edition. Last accessed May 6, 2014: 

http://des.nh.gov/organization/commissioner/pip/publications/wd/documents/wd-06- 

34.pdf. New Hampshire Department of Environmental Services. 

NHNHB. 2008. New Hampshire natural heritage bureau. Rare plants, rare animals, and 

exemplary natural communities in New Hampshire towns. Page 183 in D.-D. o. F. Lands, 

editor. 

Nowak, D. J. 2002. The effects of urban trees on air quality. United States Department of 

Agriculture, Forest Service. 

Nowak, D. J., K. L. Civerolo, S. Trivikrama Rao, S. Gopal, C. J. Luley, and D. E. Crane. 2000. 

A modeling study of the impact of urban trees on ozone. Atmospheric Environment 

34:1601-1613. 

Nowak, D. J., D. E. Crane, and J. C. Stevens. 2006. Air pollution removal by urban trees and 

shrubs in the United States. Urban Forestry and Urban Greening 4:115-123. 

Nowak, D. J., R. E. Hoehn, D. E. Crane, J. C. Stevens, and C. Leblanc Fisher. 2010. Assessing 

urban forest effects and values, Chicago’s urban forest. Resour. Bull. NRS-37. Available 

at: http://www.itreetools.org/resources/reports/Chicago's%20Urban%20Forest.pdf. U.S. 

Department of Agriculture, Forest Service, Northern Research Station, Newtown Square, 

PA. 

Nowak, D. J., R. E. Hoehn, D. E. Crane, J. C. Stevens, and J. T. Walton. 2007. Assessing urban 

forest effects and values, New York City’s urban forest. Resour. Bull. NRS-9. Available 

at: http://www.milliontreesnyc.org/download/pdf/ufore_study.pdf. U.S. Department of 

Agriculture, Forest Service, Northern Research Station, Newtown Square, PA. 

Nowak, D. J., J. E. Pasek, R. A. Sequeira, D. E. Crane, and V. C. Mastro. 2001. Potential effect 

of Anoplophora glabripennis (Coleoptera: Cerambycidae) on urban trees in the United 

States. Journal of Economic Entomology 94:116-122. 

NRC. 2007. Status of pollinators in North America. National Academy Press, Washington, DC. 

Ohio Department of Natural Resources. nd. Best Management Practices for Logging Operations: 

Factsheet. Ohio Department of Natural Resources, Division of Forestry, Columbus, Ohio. 

Oliver, C. D. and B. C. Larson. 1996. Forest Stand Dynamics. John Wiley and Sons, Inc., New 

York. 

Orwig, D. A., R. C. Cobb, A. W. D'Amato, M. L. Kizlinski, and D. R. Foster. 2008. Multi-year 

ecosystem response to hemlock woolly adelgid infestation in southern New England 

forests. Canadian Journal of Forestry Research 38:834-843. 

Orwig, D. A. and D. R. Foster. 1998. Forest response to the introduced hemlock woolly adelgid 

in southern New England, USA. Torrey Botanical Society 125:60-73. 

Pavlaki, M. D., R. Pereira, S. Loureiro, and A. M. V. M. Soares. 2011. Effects of binary mixtures 

on the life traits of Daphnia magna. Ecotoxicology and Environmental Safety 74:99-110. 

Perkins, D. W. and M. L. J. Hunter. 2006. Effects of riparian timber management on amphibians 

in Maine. Journal of Wildlife Management 70:657-670. 

Peterson, A. T., R. Scachetti-Pereira, and W. W. Hargrove. 2004. Potential geographic 

distribution of Anoplophora glabripennis (Coleoptera: Cerambycidae) in North America. 

American Midland Naturalist 151:170-178. 

Petranka, J. W. and C. K. Smith. 2005. A functional analysis of streamside habitat use by 

southern Appalachian salamanders: Implications for riparian forest management. Forest 

Ecology and Management 210:443-454. 
Appendix I. References I–11 

Pettis, J. S. and K. S. Delaplane. 2010. Coordinated responses to honey bee decline in the USA. 

Apidologie 41:256-263. 

Pettis, J. S., D. vanEngelsdorp, J. Johnson, and G. Dively. 2012. Pesticide exposure in honey 

bees results in increased levels of the gut pathogen Nosema. Naturwissenschaften 99:153- 

158. 

Phillips, P. J. and R. W. Bode. 2004. Pesticides in surface water runoff in south-eastern New 

York State, USA: Seasonal and stormflow effects on concentrations. Pest Management 

Science 60:531-543. 

Pineda, E., C. Moreno, F. Escobar, and G. Halffter. 2005. Frog, bat, and dung beetle diversity in 

the cloud forest and coffee agroecosystems of Veracruz, Mexico. Conservation Biology 

19:400-410. 

Poland, T. M., R. A. Haack, T. R. Petrice, D. L. Miller, L. S. Bauer, and R. Gao. 2006. Field 

evaluations of systemic insecticides for control of Anoplophora glabripennis (Coleoptera: 

Cerambycidae) in China. Journal of Economic Entomology 99:383-392. 

Poland, T. M. and D. G. McCullough. 2006. Emerald ash borer: Invasion of the urban forest and 

the threat to North America’s ash resource. Journal of Forestry:118-124. 

Pope III, C. A., R. T. Burnett, G. D. Thurston, M. J. Thun, E. E. Calle, D. Krewski, and J. J. 

Godleski. 2003. Cardiovascular mortality and long-term exposure to particulate air 

pollution: Epidemiological evidence of general pathophysiological pathways of disease. 

Circulation 109:71-77. 

Potts, S. G., J. C. Biesmeijer, C. Kremen, P. Neumann, O. Schweiger, and W. E. Kunin. 2010. 

Global pollinator declines: Trends, impacts and drivers. Trends in Ecology and Evolution 

26:345-353. 

Price, J. I., D. W. McCollum, and R. P. Berrens. 2010. Insect infestation and residential property 

values: A hedonic analysis of the mountain pine beetle epidemic Forest Policy and 

Economics 12:415-422. 

Randrup, T. B., E. G. McPherson, and L. R. Costello. 2001. A review of tree root conflicts with 

sidewalks, curbs, and roads. Urban Ecosystems 5:209-225. 

Richter, B. D., D. P. Braun, M. A. Mendelson, and L. L. Master. 1997. Threats to imperiled 

freshwater fauna. Conservation Biology 11:1081-1093. 

Sanchez-Bayo, F. and K. Goka. 2014. Pesticide residues and bees – A risk assessment. PLoS 

One 9:1-16. 

Sarajevs, V. 2011. Health Benefits of Street Trees. Forest Research. 

Sawyer, A. J. 2006. Infestation dynamics of the Asian long-horned beetle in the United States.in 

Proceedings 17th U. S. Department of Agriculture Interagency Research Forum on 

Gypsy Moth and other invasive species 2005. 

Scarpa, R., W. G. Hutchinson, S. M. Chilton, and J. Buongiorno. 2000. Importance of forest 

attributes in the willingness to pay for recreation: a contingent valuation study of Irish 

forests. Forest Policy and Economics 1:315-329. 

Schmuck, R., R. Schöning, A. Stork, and O. Schramel. 2001. Risk posed to honeybees (Apis 

mellifera, Hymenoptera) by an imidacloprid seed dressing of sunflowers. Pest 

Management Science 57:225-238. 

Shifley, S. R., F. X. Aguilar, N. Song, S. I. Stewart, D. J. Nowak, D. D. Gormanson, W. K. 

Moser, S. Wormstead, and E. J. Greenfield. 2012. Forests of the Northern United States. 

Gen. Tech. Rep. NRS-90. Last accessed May 29, 2014 

http://www.fs.fed.us/nrs/pubs//gtr/gtr_nrs90.pdf. United States Department of 

Agriculture, Forest Service, Northern Research Station, Newtown Square, PA. 
I–12 Appendix I. References

Simpson, J. R. 1998. Urban forest impacts on regional cooling and heating energy use: 

Sacramento County case study. Journal of Arboriculture 24. 

Smith, M. T., J. Bancroft, G. Li, R. Gao, and S. Teale. 2001. Dispersal of Anoplophora 

glabripennis (Cerambycidae). Environmental Entomology 30:1036-1040. 

Smith, M. T., J. Bancroft, and J. Tropp. 2002. Age-specific fecundity of Anoplophora 

glabripennis (Coleoptera: Cerambycidae) on three tree species infested in the United 

States. Environmental Entomology 31:76-83. 

Smith, M. T., P. C. Tobin, J. Bancroft, G. Li, and R. Gao. 2004. Dispersal and spatiotemporal 

dynamics of Asian longhorned beetle (Coleoptera: Cerambycidae) in China. 

Environmental Entomology 33:435-442. 

Smith, W. B., P. D. Miles, C. H. Perry, and S. A. Pugh. 2007. Forest Resources of the United 

States, 2007 (last accessed June 30, 2014 at 

http://www.fs.fed.us/nrs/pubs/gtr/gtr_wo78.pdf). United States Department of 

Agriculture, Forest Service. 

Smock, L. A. and C. M. MacGregor. 1988. Impact of the American chestnut blight on aquatic 

shredding macroinvertebrates. Journal of the North American Benthological Society 

7:212-221. 

Starner, K. and K. S. Goh. 2012. Detections of the neonicotinoid insecticide imidacloprid in 

surface waters of three agricultural regions of California, USA, 2010–2011. Bull. 

Environ. Contam. Toxicol. 88:316-321. 

Stern, A. C. 1977. Air Pollution: The effects of air pollution. Elsevier. 

Suarez, A. V., K. S. Pfening, and S. K. Robinson. 1997. Nesting success of a disturbancedependent songbird on different kinds of edges. Conservation Biology 11:928–935. 

Sydnor, T. D., M. Bumgardner, and A. Todd. 2007. The potential economic impacts of emerald 

ash borer (Agrilus planipennis) on Ohio, U.S., communities. Arboriculture and Urban 

Forestry 33:48-54. 

Takano, T., K. Nakamura, and M. Watanabe. 2002. Urban residential environments and senior 

citizens' longevity in megacity areas: the importance of walkable green spaces. J 

Epidemiol Community Health 56:913-918. 

Tasei, J. N. 2002. Impact of agrochemicals on non-Apis bees. Pages 101-131 in J. Devillers and 

M.-H. Pham-Delegue, editors. Honey Bees: Estimating the Environmental Impacts of 

Chemicals. CRC Press. 

Taylor, A. F., F. E. Kuo, and W. C. Sullivan. 2002. Views of nature and self-discipline: evidence 

from inner city children. J Environ Psychol 22:49-63. 

Thompson, F. R. I. 2005. Landscape level effects on forest bird populations in eastern broadleaf 

forests: Principles of conservation. Pages 290-295 in USDA-FS, editor. General 

Technical Report. 

Thompson, H. M. 2003. Behavioural effects of pesticides in bees–Their potential for use in risk 

assessment. Ecotoxicology 12:317-330. 

Thurber, D. K., W. R. McClain, and R. C. Whitmore. 1994. Indirect effects of gypsy moth 

defoliation on nest predation. Journal of Wildlife Management 58:493-500. 

Tiwary, A., D. Sinnett, C. Peachey, Z. Chalabi, S. Vardoulakis, T. Fletcher, G. Leonardi, C. 

Grundy, A. Azapagic, and T. R. Hutchings. 2009. An integrated tool to assess the role of 

new planting in PM10 capture and the human health benefits: a case study in London. 

Environmental Pollution 157:2645-2653. 
Appendix I. References I–13 

Tu, C. M. 1995. Effect of five insecticides on microbial and enzymatic activities in sandy soil. 

Journal of Environmental Science and Health, Part B: Pesticides, Food Contaminants, 

and Agricultural Wastes 30:289-306. 

Tzoulas, K., K. Korpela, S. Venn, V. Yli-Pelkonen, A. Ka ́zmierczak, J. Niemela, and P. 

James. 2007. Promoting ecosystem and human health in urban areas using Green 

Infrastructure: A literature review. Landscape and Urban Planning 81:167-178. 

Ulrich, R. S. 1984. View through a window may influence recovery from surgery. 

Science 224:420-421. 

University of Vermont Tourism Data Center. 2012. The Vermont travel and tourism industry - 

2012 (last accessed June 30, 2014 at 

http://www.uvm.edu/~snrvtdc/publications/Vermont_Tourism_2011.pdf). 

USDA. 2009. Census of Agriculture: Census of horticultural crops, 2009. Last accessed July 2, 

2014 at 

http://www.agcensus.usda.gov/Publications/2007/Online_Highlights/Census_of_Horticu

l ture_Specialties/. United States Department of Agriculture. 

USDA. 2012. Report on the National Stakeholders Conference on Honey Bee Health. United 

States Department of Agriculture, Alexandria, Virginia. 

USGS. 1998. Northeast wetland flora. Last accessed May 6, 2014: 

http://www.npwrc.usgs.gov/resource/plants/florane/list.htm. United States Geological 

Service (USGS). 

USGS. 2014. Digital representations of tree species range maps from “Atlas of United States 

Trees” by Elbert L. Little Jr. (and other publications). Last accessed July 1, 2014 at 

http://esp.cr.usgs.gov/data/little/. United States Department of the Interior, U.S. 

Geological Survey. 

Velarde, M. D., G. Fry, and M. Tveit. 2007. Health effects of viewing landscapes -Landscape 

types in environmental psychology. Urban Forestry & Urban Greening 6:199-212. 

Vidau, C., M. Diogon, J. Aufauvre, R. Fontbonne, B. Viguès, J.-L. Brunet, C. Texier, D. G. 

Biron, N. Blot, H. El Alaoui, L. P. Belzunces, and F. Delbac. 2011. Exposure to 

sublethal doses of fipronil and thiacloprid highly increases mortality of honeybees 

previously infected by Nosema ceranae. PLoS One 6:1-8. 

Wall, D. H., R. D. Bardgett, V. Behan-Pelletier, J. E. Herrick, T. H. Jones, K. Ritz, and J. Six. 

2012. Soil ecology and ecosystem services. Oxford University Press. 

Walters, R. S. and H. W. Yawney. 1990. Acer rubrum L. Red maple. Silvics of North America 

2:60. 

Wan, M. T., D. J. Mout, and R. G. Watts. 1987. Acute toxicity to juvenile pacific salmonids 

of Garlon 3ATM, Garlon 4TM, triclopyr, triclopyr ester, and their transformation 

products: 3,5,6-trichloro-2-pyridinol and 2-methoxy-3,5,6-trichloropyridine. Bull. 

Environ. Contam. Toxicol. 39:721-728. 

Wang, B., V. C. Mastro, and W. H. McLane. 2000. Impacts of chipping on surrogates for the 

longhorned beetle Anoplophora glabripennis (Coleoptera: Cerambycidae) in logs. 

Journal of Economic Entomology 93:1832-1836. 

Wear, D. N. and J. G. Greis. 2013. The Southern Forest Futures Project: Technical Report. Gen. 

Tech. Rep. SRS-178. Last accessed May 16, 2014, 

http://www.srs.fs.fed.us/pubs/gtr/gtr_srs178.pdf. U.S. Department of Agriculture, 

Forest Service, Southern Research Station, Asheville, NC. 
I–14 Appendix I. References

Wenger, S. 1999. A review of the scientific literature on riparian buffer width, extent, and 

vegetation. University of Georgia, Institute of Ecology, Office of Public Service and 

Outreach. 

Whitelaw, S., J. Swift, A. Goodwin, and D. Clark. 2008. Physical activity and mental health: 

The role of physical activity in promoting mental wellbeing and preventing mental 

health problems: An evidence briefing. University of Glasgow. 

Williams, D. W., G. Li, and R. Gao. 2004. Tracking movements of individual Anoplophora 

glabripennis (Coleoptera: Cerambycidae) adults: application of harmonic radar. 

Environmental Entomology 33:644-649. 

Woods, K. D. 1997. Community Response to Plant Invasion. Pages 56-68 Assessment and 

Management of Plant Invasions. Springer-Verlag, Inc., New York. 

Yarie, J. 1980. The role of understory vegetation in the nutrient cycle of forested ecosystems in 

the mountain hemlock biogeoclimatic zone. Ecology 61:1498-1514. 

Zhang, Y.-H., S.-S. Liu, H.-L. Liua, and Z.-Z. Liuc. 2010. Evaluation of the combined toxicity 

of 15 pesticides by uniform design. Pest Management Science 99:879-887. 
Appendix J. Distribution List J–1 

Appendix J. Distribution List

Patrick Atagi, Vice President 

Advocacy and External Affairs 

Natl. Wooden Pallet & Container Assoc. 

1421 Prince Street, Suite 340 

Alexandria, VA 22314-2805 

patagi@palletcentral.com

Phillip M. Baldauf, Project Manager 

USDA–APHIS– PPQ 

1761 State Route 125, Suite A 

Amelia, OH 45102 

phillip.m.baldauf@aphis.usda.gov

John Burch, State Plant Health Director 

USDA–APHIS–PPQ 

8995 East Main Street 

Reynoldsburg, OH 43068 

John.M.Burch@aphis.usda.gov

Clinton E. Burfitt, Program Manager 

Insect Pest Prevention & Mgmt. Program 

Plant Program Area, OR Dept. of Agric. 

635 Capitol St NE, Room 238 

Salem, OR 97301-2532 

cburfitt@oda.state.or.us

Faith T. Campbell, Ph.D. 

Vice President 

Center for Invasive 

Species Prevention 

8208 Dabney Avenue 

Springfield, VA 22152 

phytodor@aol.com

Lee Corte-Real, Director 

Division of Crop and Pest Services 

MA Department of Agricultural Resources 

251 Causeway Street, Ste. 500 

Boston, MA 02114–2151 

lee.corte-real@state.ma.us

 

 

Carol DiSalvo 

IPM Coordinator 

NPS Servicewide 

carol_disalvo@nps.gov

Timothy Estep 

151 West Boylston Dr 

Worcester, MA, 01606 

timothy.estep@aphis.usda.gov

Troy Evans 

Kodak, TN 37764 

http://www.springwildflowerpilgrimage.org/

Geir Y. Friisoe 

c/o Aurelio Posadas, Executive Secretary 

P.O. Box 847 

Elk Grove, CA 95759 

AURELIOP@ELKGROVE.NET

Brian Gallagher, Senior Director, 

Conservation Science 

Western Pennsylvania Conservancy 

800 Waterfront Drive 

Pittsburgh, PA 15222 

info@paconserve.org 

Steve Garske 

Policy Analyst/Attorney 

 Great Lakes Indian Fish & Wildlife Comm. 

72682 Maple Lane 

P.O. Box 9 Odanah, 

WI 54861 

steveg@glifwc.org

Richard Gaskalla, Division Director 

Division of Plant Industry 

 Florida Department of Agriculture & 

Consumer Services 

1911 SW 34th Street 

Gainesville, FL 32608 

gaskalr@doacs.state.fl.us
J–2 Appendix J. Distribution List

Michael A. Girard 

Executive Director 

North American Maple Syrup Council 

Simsbury, CT 06070 

mgirard@simscroft.com 

www.northamericanmaple.org

Joseph P. Gittleman, Project Manager 

ALB Eradication Program, New York 

320–01 Merrick Rd 

Amityville, NY 11701 

Joe.P.Gittleman@aphis.usda.gov

Ken Gooch 

Forest Health Program Director 

Massachusetts Bureau of Forestry 

51 Military Road 

Amherst, MA 01002 

ken.gooch@massmail.ma.us

Matthew Gordon 

c/o Vermont Maple Sugar Makers' Assn. 

Waterbury Center, VT 05676 

Email: mgordon@vermontmaple.org

Kathleen Gough 

c/o Tate Township Hall 

3001 SR 125 

Bethel, OH 45106–9454 

Steve Hobart 

Blanchard Twp, ME 04406 

Ellie Irons, Program Manager 

Commonwealth of Virginia 

Dept. of Environmental Quality 

P.O. Box 1105 

Richmond, VA 23218 

www.deq.virginia.gov

George Jibilian 

314 Peterson St. 

Fort Collins, CO 80524 

George@multigrain.com

Dan Kenny, State Plant Regulatory Official 

Plant Pest Control Section 

Division of Plant Industry 

Ohio Dept. of Agriculture 

8995 East Main St. 

Reynoldsburg, OH 43068 

dkenny@agri.ohio.gov

Elizabeth A. Knott 

9 Hillcroft Ave. 

Worcester, MA 01606 

David Lance, Acting Director 

USDA-APHIS-PPQ CPHST 

Otis Laboratory 

1398 W Truck Rd. 

Buzzards Bay, MA 02542 

david.r.lance@aphis.usda.gov

Elaine Leslie 

National Park Service 

Biological Resource Management Division 

1201 Oakridge Drive, Suite 200 

Fort Collins, CO 80525 

Christopher Logue, Director 

Division of Plant Industry 

New York Dept. of Agriculture and 

Markets 10 B Airline Drive 

Albany, NY 12235–0001 

Christopher.Logue@agriculture.ny.gov

Joan Mahoney 

Division of Plant Industry 

New York Dept. of Agriculture & Markets 

10 B Airline Drive 

Albany, NY 1223–0001 

joan.mahoney@agriculture.ny.gov

Antony Massop, Acting Project Manager 

ALB Eradication Program 

151 West Boylston Dr. 

Worcester, MA 01606 

antony.j.massop@aphis.usda.gov
Appendix J. Distribution List J–3 

Antony Massop, Acting Project Manager 

ALB Eradication Program 

151 West Boylston Dr. 

Worcester, MA 01606 

antony.j.massop@aphis.usda.gov

Barry Matthews 

Commonwealth of Virginia 

Dept of Health, Madison Building 

109 Governor Street, 6th Floor 

Richmond, VA 23219 

www.vdh.virginia.gov

Dennis Meguire 

4922 State Route 22 

Batavia, OH 45103 

Jocelyn Millar 

Department of Entomology 

University of California 

Riverside, CA 92521 

jocelyn.millar@ucr.edu

 

Dr. Timothy Perkins 

Professor & Director 

Proctor Maple Research Center 

University of Vermont 

Timothy.Perkins@uvm.edu

Winton Pitcoff 

Massachusetts Maple Producers Assoc. 

Plainfield, MA 01070 

info@massmaple.org

Robyn Rose, PhD 

National Program Manager 

USDA, APHIS, PPQ 

4700 River Road, Unit 26 

Riverdale, MD 20737 

robyn.i.rose@aphis.usda.gov

Ronald E. Rothhaas, Jr. 

P.O. Box 11094 

Cincinnati, OH 45211 

Josie Ryan, National Operations Manager 

ALB Eradication Program 

USDA, APHIS, PPQ 

320 Merrick Rd. 

Amityville, NY 11701 

josie.k.ryan@aphis.usda.go

Rhonda Santos

Public Information Officer 151 

West Boylston Drive 

Worcester, MA 01606

rhonda.J.Santos@aphis.usda.gov

Branislav Simic 

Acting State Plant Health Director 

97 Barnes Rd., Unit 200 

Wallingford, CT 06492 

brani slav.v.simic@aphis.usda.gov

Bettina Sullivan, Manager 

Environmental Impact Review and 

Long-Range Priorities 

Commonwealth of Virginia 

Department of Environmental Quality 

P.O. Box 1105 

Richmond, VA 23218 

Bettina.sullivan@deq.virginia.gov

Charles.ellis@deq.virginia.gov

Glenn Taylor 

Great Smoky Mountains National Park 

107Park Headquarters Road 

Gatlinburg, TN 37738 

glenn_taylor@nps.gov

The Nature Conservancy 

Arlington, VA 22203 

Roy Unversaw 

21444 Woodville Road 

Blanchester, OH 45107 
J–4 Appendix J. Distribution List

Wendy Walsh 

927 N. Main Street, Bld. D 

FWS – NJ Field Office Ecological Svcs. 

Pleasantville, New Jersey 08232 

http://www.fws.gov/northeast/njfieldoffice/

Malfalda Weldon 

Acting NY State Plant Health Director 

USDA–APHIS–PPQ 

Penbrooke Dr., Suite 14 

Penfield, NY 14526 

malfalda.weldon@aphis.usda.gov 
Appendix K. Index K–1 

Appendix K. Index

A

Adaptive management, 25, 85 

Additional methods development and research, 33 

AFFECTED ENVIRONMENT, 35

Affected Biological Organisms, 45 

Affected Economic, Social, and Cultural Factors,

48 

Affected Forest Communities, 35

Affected Soil, 39 

Affected Water Resources, 36 

Air, 64, 72, 76, 83 

ALB 

ALB host species and soil quality, 39 

Dispersal and Spread, 4 

Economic impacts of ALB in the United States, 7

Existing and projected ecological impacts in the 

United States and elsewhere, 9

Geographic distribution, 6 

History of actions against ALB, 9 

Hosts, 3 

Life cycle, 2 

ALTERNATIVES, 19

Alternatives considered but not included in this

EIS, 32 

Other Integrated Pest Management/ Suppression 

Strategies, 32 

Use of other chemical control agents, 32

Alternatives evaluated in this EIS, 25

Quarantine, 19 

Survey, 22 

Applicable environmental statutes, 90

Authority to act, 1 

B

Bald and Golden Eagle Protection Act, 95 

Biological control, 32 

C

Consultations, 17 

Cultural factors, 48 

Cumulative impacts, 85 

D

Decision framework, 15 

Description of ALB, 2-7 

E

Ecological resources, 58, 66, 73, 77, 84

Economic impacts of ALB in the U.S., 7 

Economic, social, and cultural resources, 60, 68, 73, 

80, 84 

Endangered Species Act, 90 

 

E, continued.

ENVIRONMENAL CONSEQUENCES, 55

Program Alternatives, 55 

Special Programmatic Considerations, 90

Environmental resources, 64, 71, 75 

Executive Order 12898, 96 

Executive Order 13045, 96 

Executive Order 13175, 97 

Executive Order 13186, 97 

EXECUTIVE SUMMARY, v

F

Federal and State Regulations and Statutes, 97

Forest resources, 55, 63, 71, 75, 83 

Full Host Removal, 26, 71 

 H

Herbicide use, 66, 70 

Human health, 62, 70, 74, 81, 85 

 I

Imidacloprid, 75, 84, 88-90 

Infested tree removals and insecticide treatment of

high-risk host trees, 75 

Insecticide treatment, 30 

Insecticide/herbicide use, 77, 81 

Integrated Approach (Preferred alternative), 26, 

83 

Interactions of soils with water quality, 39

Irreversible and Irretrievable Commitment of 

Resources, 100 

M

Maple syrup, 51 

Market factors, 48 

Migratory Bird Treaty Act, 94

Mitigation and risk reduction, 98 

N

NEPA requirements, 17 

No action, 26, 55 

Non-market factors, 51 

P

Pheromones (insect), 33 

Pollinators, 47 

Preferred Alternative, 26, 83 

Program monitoring, 100 

Public involvement, 11 
K–2 Appendix K. Index 

P, Continued 

PURPOSE OF and NEED FOR ACTION, 1

Background, 2 

Consultations, 17 

Decision Framework, 15 

Public Involvement, 11 

Scope of This Document and NEPA Requirements, 

17 

 Q

Quarantine, 21 

R

Regulatory treatments, 33 

Removal of infested trees, 26, 63, 75 

S

 Scope of document and NEPA Requirements, 17 

Social factors, 53 

Soil, 66, 72, 76, 84 

Soil Erosion, 41 

Soil types of the United States, 40

Stakeholder issues, 11 

Sterile insect technique, 32 

T

Timber, 49 

Triclopyr, 66, 70-71, 76 

Tree nurseries, 51 

 W

Water, 64, 71, 75, 83 

Water quality, 37, 39, 55, 64, 72, 89